A. Purpose

The use of oxygenated gasoline was mandated under the Clean Air Act
Amendments of 1990 in areas that did not meet the federal ambient air
standard for carbon monoxide (CO), an air pollutant associated with
potential health risks. CO interferes with the body's ability to
utilize oxygen by combining with hemoglobin, which then prevents it from
transporting oxygen efficiently to organs of the body. Persons with
chronic heart disease are at particular risk for adverse health effects
of CO, but other groups, including the elderly, pregnant women, infants,
and persons with anemia or cardiopulmonary disease, are also likely to
be at increased risk due to CO exposure. Motor vehicle emissions are
the primary source of ambient CO levels in most areas. The Clean Air
Act requires at least a 2.7% oxygen content for gasoline sold in CO
nonattainment areas. Gasoline containing 2.7% oxygen (by weight) is
typically achieved by the addition of 15% methyl tertiary butyl ether
(MTBE) or 7.5% ethanol (by volume). The higher oxygen content of
oxygenated gasoline compared to conventional gasoline is intended to
lead to a more complete combustion of the gasoline and therefore to
reduced tailpipe emissions of CO. In addition, oxygenates usually
displace aromatics in gasoline, such as benzene, toluene and xylene,
which are a source of octane. These aromatics have been targeted for
reduction in the reformulated gasoline program.

Soon after the oxygenated gasoline program was introduced
nationally in the winter of 1992-1993, anecdotal reports of acute health
symptoms were received by health authorities in various areas of the
country. Such health concerns had not been anticipated but have
subsequently focused attention on the possible health risks associated
with using oxygenated gasoline. The purpose of this interagency
assessment of the potential health risks of oxygenated gasoline is to
explore the question of whether evidence from recent health studies of
oxygenated gasoline or oxygenates, especially MTBE, and acute illness
warrants a reconsideration by EPA of potential health risks of the
oxygenated gasoline program during the coming winter months. This
assessment is being prepared for the Office of Science and Technology
Policy, under the auspices of the Interagency Oxygenated Fuels
Assessment Steering Committee. The potential health benefits of the
oxygenated program are not addressed in this assessment; such
benefits will be considered in the more comprehensive interagency
assessment of oxygenated fuels that is currently under way.

A team of scientists from three Federal agencies, the Centers for
Disease Control and Prevention (CDC), the National Institute for
Environmental Health Sciences (NIEHS), and the Environmental Protection
Agency (EPA) was assembled to complete this evaluation. Members of this
team were chosen by members of the Interagency Steering Committee or
their designates on the basis of the members' expertise in different
scientific disciplines and knowledge about issues related to oxygenated
gasoline.

B. Scope

The interagency team assembled for this short-term assessment was
asked to review studies that have been made available since the
enactment of the Clean Air Act Amendments of 1990. Because much of this
recent research has not been published, the team evaluated unpublished
as well as published reports. Because of the limitations of available
data, this assessment focuses primarily on MTBE in oxygenated gasoline
and its principal metabolite, tertiary butyl alcohol (TBA). Other
ethers and alcohols, including ethanol, ethyl tertiary butyl ether
(ETBE), tertiary amyl methyl ether (TAME), tertiary amyl ethyl ether
(TAEE), and diisopropyl ether (DIPE), are less extensively used (or may
potentially be used) in oxygenated gasoline. Although this assessment
does not address these other oxygenates as extensively as MTBE, this
does not imply that MTBE is the only oxygenate of concern. The focus of
this assessment is oxygenated gasoline, not reformulated gasoline.
Reformulated gasoline is intended to reduce motor vehicle emissions of
hydrocarbons which lead to higher ozone (smog) levels during summer
months and air toxics year round. Reformulated gasoline has 2.0% oxygen
(by weight), typically achieved by the addition of 11% MTBE or 5%
ethanol (by volume), as well as lower concentrations of certain volatile
organic compounds than conventional gasoline.

This assessment focuses on exposures that occur through inhalation;
ingestion of oxygenates from contaminated water supplies may also be
important but was considered beyond the scope of this assessment. The
data base on ethanol is extensive, but it pertains primarily to
ingestion, not inhalation. A full consideration of the relevance of the
literature regarding the health effects of ethanol by ingestion to
exposure by inhalation from evaporative and combustion mixtures of
ethanol in gasoline also was considered beyond the scope of this
assessment.

This assessment also attempts to identify areas where the scientific
data base is particularly limited. It should also be noted that most of
the information available on the health effects of MTBE pertains to MTBE
alone rather than the mixture of MTBE and gasoline. This assessment
does not attempt to assess the risks or benefits of MTBE-oxygenated
gasoline in relation to conventional gasoline.

In addition, a more comprehensive review of the fundamental
basis and efficacy of the Environmental Protection Agency's winter
oxygenated gasoline program is currently being conducted by a
combination of technical and scientific experts from across several
Federal agencies, under the coordination of the National Science and
Technology Council's Committee on Environment and Natural Resources.
This more comprehensive review will consider not only health effects,
but also air quality benefits, fuel economy and engine performance,
ground water and drinking water quality, and an economic analysis of
benefits. In addition, an independent review of the health effects of
the use of oxygenates in gasoline is currently being conducted by the
Health Effects Institute (HEI) and a panel of experts, in response to a
request from the EPA. It is expected that the HEI review will be the
core of the health effects section of the more comprehensive interagency
review.

A. Air Quality

Various studies (e.g., Stump et al., 1989, 1990, 1992, 1994;
Auto/Oil Air Quality Improvement Research Program 1990, 1991a,b,c;
Carter et al., 1991; Kiskis et al., 1989; Reuter et al., 1992) have
examined the impact of oxygenates on evaporative and combustion
emissions from vehicles. For example, one study (Reuter et al., 1992)
of 20 vehicles (1989 models with three-way catalysts) showed that, at
24°C, there was a trend for MTBE to cause a net reduction in the total
mass of toxic combustion emissions. In general, emissions of benzene
tended to decrease and emissions of formaldehyde tended to increase;
1,3-butadiene was not substantially altered when MTBE was added.
However, more testing is needed with different types of vehicles under a
variety of conditions (especially lower and higher temperatures) to
characterize fully the overall changes in evaporative and combustion
emissions associated with oxygenate usage.

Although some models for relating emissions data to ambient air quality
are available, these changes in emissions cannot be quantitatively
extrapolated to estimate with acceptable precision their impact on the
air quality of a city. Moreover, ambient air measurements may provide
only a rough (and sometimes misleading) guide to personal exposure
levels that a population might encounter during hour-to-hour and
day-to-day activities.

Little is known about the atmospheric transport and transformation
of the complex mixtures associated with the use of oxygenated fuels or
the possible significance of human exposure to these fuels. For
example, tertiary butyl formate is a photochemical-transformation
product of MTBE, as is formaldehyde (Tuazon et al., 1991). The
possibility of increased human exposures to such substances as a result
of using oxygenated fuel has not yet been evaluated. Ambient and
microenvironmental air monitoring is needed to determine the levels of
these and other possibly significant by-products of oxygenated fuels.

1. Ambient air measurements

Zweidinger (1993) analyzed air samples collected over 8-hour
periods in Fairbanks, Alaska; Stamford, Connecticut; and Albany, New
York for aldehydes, MTBE, and other volatile organic compounds (VOCs) in
conjunction with investigations by CDC (1993a,b,c). The Fairbanks
samples were collected by the State of Alaska during three time periods:
time period 1 (December 1 through 12, 1992) occurred immediately before
the phase out of 15% MTBE in gasoline, and 25 VOC and 35 aldehyde
samples were obtained; time period 2 occurred during the phase-out
itself (December 18 through 22, 1992), and 31 VOC and 26 aldehyde
samples were obtained; and time period 3 occurred after the phase-out of
MTBE-oxygenated fuels (February 2 through March 5, 1993) during which
73 samples each of VOCs and aldehydes were obtained. On the basis of
the analysis of gasoline samples collected in Fairbanks during time
periods 2 and 3, the percentage of MTBE in gasoline decreased from 8.5%
to 1% for unleaded regular gasoline and from 14.7% to 5.6% for premium
gasoline. The Stamford samples were collected by EPA Region 1 from
April 13 through 14, 1993; 30 samples each of VOCs and aldehydes were
collected. These samples were obtained in Stamford because the city
represented another part of the country where MTBE-oxygenated gasoline
was also sold. The Albany samples were collected from May 5 through May
27, 1993, by the New York State Department of Health; 20 samples each of
VOCs and aldehydes were collected. Albany represented an area of the
country where MTBE was not used as an oxygenate but was present only as
an octane enhancer in gasoline. The Fairbanks time-period 1 VOC samples
were analyzed by the Oregon Graduate Center, and the Desert Research
Institute analyzed the aldehyde samples. All other ambient samples
were analyzed by EPA's Atmospheric Research and Exposure Assessment
Laboratory.

The samples from each city consisted of samples taken at roadside
intersections, the pump islands of gas stations, garage service bays,
and in residential neighborhoods. Samples of indoor air and from
background sites were also collected. Samples of indoor air and of air
from service bay or background sites were not collected in Albany, and
no air samples from service bays were collected in Fairbanks during time
period 2. In addition, air samples from the interiors of commercial
vehicles in Fairbanks were collected early during time periods 1 and 3.
Significant differences in ambient temperature and other meteorological
conditions existed among the cities where samples were collected.
Relatively few samples were collected in a given area, and the samples
were collected over a few days only. Therefore, the data cannot be used
to describe the air quality quantitatively in any of these cities.
Rather, the data can be used to estimate approximate ranges of air
quality in the locations sampled.

The highest average concentrations of MTBE (0.345 parts per million
[ppm], or 1.28 mg/m³), benzene (0.629 mg/m³), total
nonmethane organic carbon (80.5 ppm C), and formaldehyde (0.038
mg/m³) were found in garage
service bays. One of the highest average concentrations for a single
compound was found to be 1,1,1-trichloroethane (methyl chloroform),
which exceeded 38.0 mg/m³ in service bays (Fairbanks, time period
3). Aside from the service bays, gas stations showed next highest
concentrations of MTBE (Fairbanks: 0.05 ppm, or 0.194 mg/m³,
time period 1; 0.037 ppm or 0.134 mg/m³, time period 2; and 0.006
ppm, or 0.020 mg/m³, time period 3). MTBE concentrations in
Stamford gas stations were the lowest (0.004 ppm or 0.013
mg/m³) but were likely the result of sampler location (the Albany
average was 0.024 ppm, or 0.086 mg/m³). Whereas the air
samplers in
Fairbanks and Albany were located on the pump islands, the samplers in
Stamford were located at least 15 feet from the islands. In Fairbanks,
indoor and outdoor MTBE concentrations were similar and averaged about
0.007 ppm (0.025 mg/m³) for the samples from time period 2,
falling to 0.001 ppm (0.0037 mg/m³) in time period 3, with
the
exception of one home where the average indoor value was 0.02 ppm
(0.072 mg/m³). This home had an attached garage, and air samples
taken there showed elevated levels of benzene (0.138 mg/m³) and
other compounds associated with gasoline. Indoor MTBE concentrations in
Stamford averaged 0.0006 ppm (0.002 mg/m³). MTBE
concentrations measured inside vehicles in Fairbanks averaged 0.007 ppm
(0.024 mg/m³) during time period 1 (not including one
sample of
0.067 ppm or 0.241 mg/m³) and averaged 0.005 ppm (0.019
mg/m³) during time period 3 (not including one sample of
0.035 ppm or 0.127 mg/m³).

Formaldehyde concentrations were higher indoors (0.012 to 0.034
mg/m³) than outdoors (0.0025 to 0.025 mg/m³), which is generally the
case, and levels appeared typical of those seen in indoor air studies.
Benzene levels were higher in Fairbanks (average roadside levels were
0.026 mg/m³ for December and 0.042 mg/m³ during time period 3)
than in the other cities (Stamford, 0.003 mg/m³ Albany, 0.0014
mg/m³). A related report by Gordian et al. (1995) noted that the
gasoline used in Alaska has the highest concentration of benzene
(~ 5%) of
any state in the nation; the report also stated that, during this same
period in Fairbanks, formaldehyde and benzene levels in indoor and
ambient air samples were higher after MTBE was removed from gasoline.
However, these findings do not establish a causal relationship between
decreased MTBE levels in gasoline and increased formaldehyde and benzene
levels in the air.

In response to complaints by residents of Milwaukee, Wisconsin,
regarding reformulated gasoline (2.0% oxygen content), the State of
Wisconsin (Anderson et al., 1995) recently measured air concentrations
of MTBE and other compounds (ETBE, benzene, ethyl benzene, xylenes,
toluene). Levels of MTBE were less than 0.001 ppm at most of the sites
sampled, including at the North Campus of the University of
Wisconsin-Milwaukee where 24-hour and 2-hour samples were obtained at
various intersections and gasoline stations. The highest ambient air
concentrations of MTBE were obtained at two gasoline stations and a
civic center parking ramp (0.00243, 0.00458, and 0.00205 ppm,
respectively). The highest value was found at a station without a stage
II vapor-recovery system.

2. Air concentrations in vehicle-related
microenvironments

The Environmental and Occupational Health Sciences Institute
(EOHSI) and the Research Triangle Institute (RTI) completed a study of
field measurements of MTBE concentrations inside automobiles during an
approximately 1-hour commute and during refueling (Lioy et al., 1993,
1994). Field measurements were collected during April 1993 in Middlesex
County, New Jersey, at two stations with full service and stage II
vapor-recovery systems; in Westchester County, New York, at three
stations with self service and stage II vapor-recovery systems; and in
Fairfield County, Connecticut, at five stations with self service and no
stage II vapor-recovery system. One new-model automobile (a 1992
Corsica) and one older model automobile, either a 1985 Caprice or a
1986 Monte Carlo, were assigned to each commuter route. The samples
were collected from the interior of the front passenger side of the
automobile. The number of samples per automobile ranged from 14 to 20
for the commute and from 3 to 5 for the fill-up.

The driver's window was open during the fill-up. The time to complete
the fill-up was about 2 minutes, and the total time at the gas station
was 5 to 10 minutes. In addition to the measurements taken inside the
automobile, a few measurements were collected near the breathing zone of
the person refueling the gas tank.

Concentrations of MTBE measured in the cabin interior during the
1-hour commute had a geometric mean of 0.006 ppm (0.021 mg/m³), with a
range of 0.001 ppm (0.004 mg/m³) to 0.16 ppm (0.58 mg/m³). The
commuter runs in Connecticut had higher geometric mean concentrations
(0.007 ppm or 0.023 mg/m³) than those in New Jersey (0.005 ppm or 0.016
mg/m³). One vehicle, the 1987 Caprice, had somewhat higher levels than
the other vehicles; inside the older-model automobiles, concentrations
were higher, probably reflecting differences in the design and
deterioration of the older-model vehicles.

Concentrations during a 5-minute refueling at a self-service station
averaged in excess of 0.3 ppm (1 mg/m³) and ranged from about 0.01 ppm
to 4.1 ppm (0.036 to 14.7 mg/m³), with the highest values measured in
the breathing zone of a person refueling at a station with no stage II
vapor-recovery system. Measurements taken inside the cars during
refueling generally ranged from 0.01 to 0.1 ppm (0.036 to 0.36 mg/m³)
for stations both with and without stage II vapor-recovery systems.

International Technologies Inc. completed a set of field
measurements of MTBE concentrations in the personal breathing zone (PBZ)
during fill-up, at the pump island, and around the property line of gas
stations (Johnson, 1993). This study was done in conjunction with the
above EOHSI/RTI study that was conducted at the same 10 gas stations.
All concentrations measured for this study, even those for intermittent
exposures in the personal breathing zone, were from a 4-hour continuous
sample. Average fence-line concentrations, which are typically taken at
the apparent property line, ranged from 0.005 to 0.065 ppm (0.018 to
0.234 mg/m³) MTBE (Johnson, 1993). The highest fence-line
concentrations ranged from 0.1 to 0.139 ppm (0.36 to 0.5 mg/m³) MTBE.
The highest breathing-zone and pump-island concentrations ranged from
0.194 to 2.5 ppm (0.7 to 9 mg/m³) MTBE. Further discussion of this
study and other exposure measurements among service station attendants
is included in the next section on occupational exposures.

As should be expected, these 4-hour breathing-zone concentrations
described above were lower than those reported by the Clayton
Environmental Consultant study (Clayton Environmental Consultants,
1991), which collected samples only during the fill-up period
(approximately 2 minutes). In the Clayton study, mean MTBE
concentrations in the breathing zone for oxygenated fuels containing 12%
to 13% MTBE were 3.6 ppm (13 mg/m³), with vapor recovery, and 8.3 ppm
(30 mg/m³), without vapor recovery. The absolute range among these
MTBE concentrations was 0.089 to 38 ppm (0.32 to 137 mg/m³). Although
several stations were monitored, the highest and lowest measurements
were made at one station, illustrating the variability of breathing zone
exposures. Indeed, a wide range of air concentrations within the
breathing zone can be expected. Ambient air concentrations measured at
a gas station will be highly dependent on wind speed and direction. In
addition, breathing-zone concentrations can be dramatically influenced
by one's position in relation to wind direction. Any spilling of fuel
while the tank is being filled also can dramatically increase the
inhaled concentration.

Personal breathing zone (PBZ) measurements obtained in conjunction with
the Wisconsin study of reformulated gasoline (2.0% oxygen content)
ranged from 0.070 to 2.93 ppm for MTBE, 0.050 to 0.100 ppm for ETBE, and
0.050 to 0.150 ppm for benzene (Anderson et al., 1995). These samples
were collected approximately 1 meter from the fuel nozzle over 15-minute
periods. The highest air concentrations of MTBE were measured at a
station lacking a Phase II vapor recovery system. In some cases, the
concentrations were higher for MTBE at a station using ETBE-oxygenated
fuel and were likely due to vapors from the previous fill-up escaping
during the refueling. Thus, it may be difficult to ascertain the
particular oxygenate to which people are exposed under real-world conditions.

B. Human Exposure Estimates for MTBE

The data on air quality and microenvironments (e.g., at gas
pumps, inside cars, in personal garages) are too limited to provide a
distribution of MTBE levels for the general population. These data,
however, may be used to roughly estimate reasonable worst-case
potential exposures, based on certain assumed activity patterns and
approximate microenvironmental concentrations. Because of the interest
in MTBE, the present evaluation focuses on this compound, even though
any potential health effects might result from complex pollutant
mixtures of which MTBE is only one component.

Acute as well as long-term exposures are of importance in
evaluating potential health risks of oxyfuels. Table 1 illustrates two
hypothetical types of people with varying activity patterns and oxyfuel
exposure potentials. The concentrations and activity patterns were
previously used by Huber (1993) for estimating MTBE exposures and were
based on a variety of population-activity studies and microenvironmental
measurements of MTBE as described above. Exposure (ppm.hr) is a
function of concentration (ppm) and time (hr). The concentrations in
Table 1 are based on measurement averages that were rounded up to the
next order of magnitude so as to provide inherent conservatism (i.e.,
overestimation) in the exposure estimates. The two estimates in Table 1
reflect different assumptions about the microenvironmental
concentrations and activity patterns represented. For example, the
concentrations of MTBE during refueling in Scenarios I and II differ by
an order of magnitude. Scenarios I and II each represent hypothetical
persons who visit a gasoline station 1.5 times per week, commute an
average of 10 hour/week, visit an auto repair shop four times per year
for 15 minutes per visit, and spend about 57 hours per week in an office
or public building. However, scenario I assumes that the person lives
in a house without an attached garage and does not reside near or spend
time around a highway or in the vicinity of a filling station.

Hypothetical person II differs from person I in that the individual
lives in a house with an attached garage and spends time outside in the
vicinity of a gasoline station or heavily used highway. The higher
concentrations of MTBE in the residential garage and home assume
evaporative emissions from the vehicle in the garage or a small gasoline
spill with the garage door closed. Hypothetical person II represents a
"reasonable worst case" scenario (Huber, 1993). More extreme cases are
conceivable, but they would probably reflect differences in the
occurrence or in the duration of activities rather than reflecting
higher average concentrations of MTBE in the microenvironments
considered. For example, a person who visits a gasoline station every
day to refuel his or her car and who works 40 hours per week outside
near a highway or other source of high MTBE levels would not likely be
exposed to higher concentrations of MTBE within those microenvironments,
but that person's overall exposure might be higher than that of
hypothetical person II because of the greater cumulative duration of his
or her exposure. Currently it is not possible to state what percentages
of the general population might be represented by hypothetical persons I
and II, but the term "reasonable worst case" is meant to imply that few
individuals would be exposed at higher levels. It is not meant to imply
a "high average" exposure. The average exposure for the entire exposed
population would be expected to be somewhat lower than hypothetical
person I and much lower than hypothetical person II.

Time-weighted average exposure levels may be estimated for the
oxyfuel season itself or for an entire year, assuming that the oxyfuel
season is limited to a few months in an area that is not required to use
reformulated gasoline. During the oxyfuel season, the time-weighted
average MTBE exposure of a hypothetical person would be one-half the sum
of that person's exposure levels divided by one-half the number of hours
in a year e.g. (159/2 ÷ 8760/2 = 0.018 ppm for person I, or 309/2 ÷
8760/2 = 0.0345 ppm for person II). To calculate an annual
time-weighted average exposure, an allowance is made for the assumed use
of 1.5% MTBE in nonoxygenated gasoline. This assumption may be a high
estimate of the average amount of MTBE in nonoxygenated gasoline
because of the relatively limited use of premium (high-octane) fuels
containing MTBE. Nevertheless, assuming 1.5% MTBE (which is 1/10 of the
15% MTBE concentration typical of oxygenated gasoline) during the
non-oxyfuel season and assuming a 6-month oxyfuel season, the annual
time-weighted average MTBE exposure level for hypothetical person I
would be [(159/8760)*(6/12)] + (0.1*[(159/8760)*(6/12)]) = 0.010 ppm.
The annual time-weighted average MTBE exposure level of hypothetical
person II would be 0.019 ppm. For a 4-month oxyfuel season, the annual
time-weighted average levels would be 0.007 ppm for person I and 0.014
ppm for person II. If reformulated gasoline containing 10% MTBE by
volume were used during the remainder of the year, the annual
time-weighted average exposure for a 6-month oxyfuel season would be
0.015 ppm for person I and 0.029 ppm for person II.

It can be assumed that a gasoline fill-up scenario, although brief,
would result in the highest acute-exposure concentrations. The highest
human exposure is expected when one is near evaporative emissions.
Thus, exposure would be greatest when handling gasoline. The highest
reported MTBE concentration measured at a filling station was 38 ppm
(137 mg/m³), although levels as low as 0.089 ppm (0.32 mg/m³) were
also measured at the same station, illustrating the variability in
fill-up exposures (Clayton Environmental Consultants, 1991). A more
typical worst-case scenario for the MTBE concentration in the breathing
zone during fill-up would be 10 ppm (36 mg/m³) MTBE for a few minutes
(Johnson, 1993; Lioy et al., 1993, 1994; Clayton Environmental
Consultants, 1991). However, higher concentrations are possible,
especially in the event of an accidental spill.

For purposes of comparison with the 1-hour human experimental exposure
studies (at 1.4 and 1.7 ppm [5 and 6 mg/m³]), discussed below, 1-hour
time-weighted average concentrations of MTBE were calculated for two
exposure scenarios by using high concentration data. The first scenario
assumed highest measured values and involved a 2-minute fill-up (38 ppm
or 137 mg/m³), a 30-minute commute associated with a fill-up (0.5 ppm or
1.8 mg/m³), and a 28-minute commute (0.076 ppm or 0.275 mg/m³); the
average MTBE concentration was 1.56 ppm (5.6 mg/m³) MTBE. The second
scenario used the MTBE levels in Table 1. Scenario 2 assumed a 2-minute
fill-up (10 ppm or 36 mg/m³), 2 minutes in a personal garage (1 ppm or
3.6 mg/m³), a 30-minute commute (0.1 ppm or 0.36 mg/m³), 10 minutes
in a public garage (0.5 ppm or 1.8 mg/m³), and 16 minutes in a public
building (0.01 ppm or 0.036 mg/m³); the average concentration was 0.5
ppm (1.8 mg/m³) MTBE.

In summary, selected hypothetical scenarios suggest that the annual
time-weighted average daily personal exposure level for a "reasonable
worst case" motorist in the general (nonoccupational) population might
be on the order of 0.019 ppm MTBE. This level is not simply a "high
average" exposure. The annual time-weighted average exposure levels for
most people would be somewhat lower, perhaps closer to the estimate of
0.010 ppm MTBE for hypothetical person I, or even an order of magnitude
lower than 0.010 ppm, in view of the rounding up of the data on which
the estimates were based. In addition, long-term exposure levels are
likely to be lower than those found during short-term measurements.
That is, hypothetical person II is unlikely to have consistently high
exposure levels from day to day, and the trend over a year or a lifetime
would be for exposure levels to regress toward the mean. Thus, in
estimating health risks in relation to these exposure estimates, it is
important to consider the likely period of exposure over which health
effects could be induced. For example, a 4 or 6-month oxyfuel season
might in itself pose little risk with respect to lifetime cancer risk
but could be of significance with respect to potential effects due to acute
exposures. Two illustrative acute-exposure estimates for different
scenarios yield averages of 0.5 ppm and 1.56 ppm, reflecting in part the
highly variable air concentrations that have been measured at gasoline
stations. It must be emphasized that these exposure values are not
based on, and make no predictions regarding, the nature of
population-exposure distributions. Personal exposure data for
probabilistic samples are needed before an exposure assessment for MTBE
can provide an adequate characterization of population exposures for
health risk assessment purposes.

TABLE 1. METHYL TERTIARY BUTYL ETHER EXPOSURE ESTIMATES

Scenario I

Scenario
II

Activity

Occurrence

Concentrationa(ppm)

Time/Year(hr)

Exposure(ppm x
hr)

Concentrationa(ppm)

Time/Year(hr)

Exposure(ppm x hr)

Refueling

1.5/week@2 min

1

2.6

2.6

10

2.6

26

Gas station

1.5/week@10 min

1

13

13

1

13

13

In vehicle

10 hr/week

0.1

520

52

0.1

520

52

Auto repair shop

4/year@15 min

0.5

1

0.5

0.5

1

0.5

Public garage

10 min/day

0.5

60

30

0.5

60

30

Residential garage

2 min/day

0.005

12

0.06

1.0

12

12

Residence

10 hr/day + weekend

0.005

4,147

20.7

0.01

4,147

41.5

Office/public buildings

57 hr/week

0.01

2,964

29.6

0.01

2,964

29.6

Outdoors

20 hr/week

0.01

1,040

10.4

0.1

1,040

104

TOTALS

8,760

159

8,760

309

Time-weighted average exposure during an Oxyfuel
Season

0.018

0.035

Annual time-weighted average exposure, assuming a 6-month
oxyfuel season

0.010

0.019

Annual time-weighted average exposure, assuming a 4-month
oxyfuel season

0.007

0.014

a Concentrations are based on measurement averages that were
rounded up to the next order of magnitude.

C. Occupational Exposure - MTBE

Since 1990, occupational exposure to MTBE has been assessed
under a variety of conditions and for various purposes, most notably in
support of evaluations of health complaints associated with exposures.
The realm of occupational exposures to MTBE involving gasoline includes
exposures to the pure compound during manufacture and transport (for
subsequent blending) and exposure to gasoline mixtures during blending,
transport, distribution, and sale. Other tangential occupational
exposures have been evaluated in jobs where at least a portion of a work
shift is spent near an MTBE-blended gasoline source; the most common
source is the automobile.

1. Service station attendants

Undoubtedly, the largest occupational group with a significant
potential exposure to MTBE is service station attendants. The number of
retail automotive service stations has been estimated at 150,000 to
210,000 (ENVIRON, 1990). Even with self service, fuel dispensing
undoubtedly continues to be an important source of occupational exposure
to MTBE, although no data were available on the number of attendants
exposed to reformulated or oxygenated fuels containing MTBE.

In 1990, at the request of the American Petroleum Institute
(API), the National Institute for Occupational Safety and Health (NIOSH)
evaluated exposure to the gasoline components MTBE, benzene, toluene,
and xylene at six retail automotive service stations (NIOSH, 1993a). To
reflect MTBE's multiple uses and the range of potential exposures, two
facilities were selected to represent its ubiquitous use as an octane
enhancer (generally blended at less than 1% of the fuel); two facilities
were selected to represent requirements to use MTBE in oxygenated fuel
(blended at 12-15 liquid volume percent [LV%] of the fuel), and two with
stage II-type vapor-recovery systems were selected to determine the
relative effectiveness of these engineering controls.

Of 16 PBZ samples collected from attendants working at stations that
sold gasoline with less than 1% MTBE, 15 were below the lowest
detectable concentration (LDC) of 0.02 ppm. The one sample above the
LDC was reported at 0.16 ppm. At stations using at least 12% MTBE
blends, 41 PBZ samples ranged from 0.03 to 3.89 ppm, averaging 0.54. At
stations equipped with stage II vapor recovery systems, 15 of the 48 PBZ
samples had detectable levels (above the LDC), ranging from 0.02 to 0.73
ppm, averaging 0.18 ppm. Sampling and analysis for MTBE was by the
NIOSH Method 1615 (Palassis, 1993). Benzene exposures among this same
population averaged approximately 0.07 ppm.

The study included an assessment of factors, such as climatic
conditions and work practices, affecting the extent of exposure.
Statistical analysis, using step-wise linear regression, indicated that
MTBE exposures were most affected by wind velocity, followed by the
amount of fuel dispensed by the attendant. The report also noted that
benzene exposures were not significantly affected by the amount of MTBE
in the fuel.

In 1994, NIOSH conducted a follow-up study to assess short-term,
or "peak," exposures among attendants (Cook, 1995). Two service
stations in New Jersey were selected, and direct-reading real-time
measurements for total hydrocarbons, as a surrogate for peak MTBE
exposures, were made in attendants' breathing zones, as were long-term,
full-shift determinations of MTBE exposures. A video-monitoring overlay
technique was used to determine the predominant sources of peak
exposures. Twenty-one of the full-shift samples taken from the PBZs of
the station attendants who were evaluated for MTBE exposure ranged from
0.08 to 1.27 ppm, with a geometric mean (GM) of 0.38 ppm. Total
hydrocarbon exposures averaged 1.89 ppm, with peaks as high as 327 ppm.
NIOSH is currently assessing the magnitude of associated MTBE peak
exposures.

Service station attendants' exposures were again assessed by NIOSH
(NIOSH, 1993b) as part of a larger effort in support of a CDC
epidemiologic investigation of public health complaints reportedly
associated with exposure to MTBE in Fairbanks, Alaska. Two samples were
obtained from attendants at service stations using MTBE as an octane
enhancer (approximately 1% MTBE). The use of MTBE as an oxygenated-fuel
had previously been discontinued after reports surfaced of ill health
effects associated with its use. MTBE concentrations were reported at
0.03 ppm (the LDC) and 0.15 ppm; both results were within the range of
exposure concentrations determined from the previous NIOSH study.

In 1993, in support of an EPA solicitation for MTBE exposure
data, the API contracted for a service station study with ITAir Quality
Services (API, 1994) at 10 service stations in the northeastern United
States. In addition to measuring MTBE, ITAQS measured benzene, toluene,
ethyl benzene, xylene, formaldehyde, and carbon monoxide in attendant
and customer breathing zones, near the pump islands, and at perimeter
locations. For comparison of sampling or analytical methods, attendant
exposures were measured using both charcoal-tube samplers and
canisters. Service stations with and without vapor-recovery systems
were included in the assessment. The MTBE content of the gasoline
ranged from 13.4 to 15.7 LV% at all stations monitored.
Of the eight, 4-hour, side-by-side charcoal tube or canister samples
collected at vapor-recovery stations, seven were above the analytical
detection limit for MTBE (detection limit not specified), ranging from
0.084 to 0.558 ppm and averaging (geometric mean) 0.33 ppm. Two 8-hour
PBZ charcoal-tube samples collected at stations without vapor-recovery
systems had MTBE concentrations of 0.554 and 1.191 ppm, respectively.

In 1994, the API contracted with NATLSCO for an evaluation of
exposure to oxygenated fuel components among attendants and mechanics at
sixteen service stations in four geographical areas (API, 1995). For
comparison purposes, sampling was conducted during the winter (February
- April) oxyfuel season, and summer (July - August) non-oxyfuel season.
During the winter, MTBE was present in the fuel of the stations in
approximate amounts ranging from 10-17 wt%. During the summer, no
oxygenates were detected above the 0.1 wt% analytical limit of detection.

During the "winter phase", 51 long-term (generally greater than six
hours) PBZ samples for MTBE ranged from 0.03 to 0.5 ppm, with a
geometric mean of 0.2 ppm. Fifty-nine short-term (generally 15 - 20
minutes) PBZ samples ranged from 0.32 - 2.1 ppm, with a geometric mean
of 0.6 ppm. During the "summer phase", 53 long-term PBZ samples ranged
from 0.03 - 0.42 ppm, with a geometric mean of 0.08. Sixty-one
short-term samples ranged from 0.19 - 0.33 ppm, averaging 0.31 ppm. The
author notes that summary statistics (range and means) include
non-detected results; they were assigned their LDC value and included in
range and GM computations. This conservative approach yields
overestimates of actual average exposures.

In 1992, the API surveyed member companies to obtain MTBE
occupational exposure data (API, 1994). Of the samples collected at
service stations, MTBE concentrations of 13 full-shift samples ranged
from 0.09 to 34 ppm, averaging 0.77 ppm (geometric mean). The author
noted that, in addition to fuel attendants, samples were collected from
mechanics, weights and measures inspectors, and people repairing fuel pumps.

2. Auto mechanics

As with service station attendants, auto mechanics and auto
technicians represent a large population with potential for significant
exposure to gasoline. In the API 16 service station study (API, 1995a),
86 long-term (generally greater than six hours),
PBZ samples for MTBE ranged from 0.02 to 2.6 ppm, averaging (geometric
mean) 0.12 ppm for the "winter phase", and from 0.02 to 0.18 ppm,
averaging 0.03 ppm during the "summer phase". Short-term samples (n=88)
ranged up to 32 ppm, with winter and summer geometric mean exposures of
1.04 and 0.42 ppm, respectively. As with the attendant sample results,
these data include non-detectable values, which overestimates the
average exposures.

In the Fairbanks study (NIOSH, 1993b), 17 of 26 PBZ samples from auto
mechanics were above the LDC for MTBE (0.03 ppm). Results ranged from
less than 0.03 to 0.45 ppm, averaging 0.06 ppm (geometric mean). As
previously mentioned, at the time of the NIOSH investigation, MTBE use
as an oxygenate had been discontinued; its use was only as an octane
enhancer (generally less than 1% MTBE/gasoline blend).

In a similar study in Stamford, Connecticut, (NIOSH, 1993c), 23 of
28 PBZ samples collected from auto mechanics were above the LDC, ranging
from less than 0.03 to 12.04 ppm, averaging 0.11 ppm. The MTBE content
of five bulk gasoline samples collected from the area ranged from 13 to 17 LV%.

In a third NIOSH study supporting a CDC epidemiologic investigation
in Albany, New York, an area not using oxygenated gasoline (NIOSH,
1993d), three of eight PBZ samples from mechanics were above the LDC,
ranging from less than 0.03 to 0.14 ppm and averaging 0.03 (geometric
mean). The MTBE content of the gasoline was reported to be in
concentrations of up to 10%.

3. Other occupational MTBE exposures related to gasoline

In the three NIOSH evaluations of MTBE exposure in support of CDC
epidemiologic investigations (NIOSH 1993b, NIOSH 1993c, NIOSH 1993d),
researchers evaluated occupations in which people had significant
exposure to motor vehicles and traffic (but not gasoline) in an effort
to determine the extent of exposure to MTBE. These occupations included
1) service station cashiers or managers, 2) service advisors, 3) parking
meter attendants, 4) animal control personnel, 5) truck drivers, and 6)
taxi drivers. Of the 26 PBZ samples collected from people in these
occupations in Fairbanks, Stamford, and Albany, only one sample,
reported at 0.10 ppm from a service station manager, was above the LDC
for MTBE. From these same investigations in Fairbanks and Stamford,
exposure to MTBE by commuters was assessed. Commuters were defined as
people whose occupations required them to spend a significant amount of
their time driving. None of 14 PBZ samples from commuters was above the
0.03 ppm LDC. The last major category assessed in these studies was
parking lot attendants. Of six PBZ samples from people in this group,
one was reported at 0.10 ppm, which is above the LDC for MTBE.

4. Other industry exposure assessments

From a survey of 17 member company's occupational exposure data for
MTBE, the API published summary statistics for PBZ long-term and
short-term samples (API, 1995b), see Table 2. Although additional
environmental data were presented in the API report, the data in Table 2
represent samples for which the sampling time was known/available. As
expected, due to the numerous contributors, sampling and analyses varied
between the OSHA Method 7, NIOSH method 1615, and in-house,
industry-developed methods.

Time-weighted average MTBE exposure related to automobiles (fuel
containing MTBE) ranged up to 0.1 ppm (only one of 26 samples was above
the 0.03 ppm LDC).

Time-weighted average exposure to MTBE associated with its
manufacture/distribution may range up to approximately 700 ppm
(transport of pure MTBE; condition not specified). The geometric mean
for this category was 0.30 ppm. The highest geometric mean for any
category was 1.73 ppm (blending pure MTBE).

D. Metabolism, Disposition, and Toxicokinetics of MTBE in Animals

The major metabolic pathway of MTBE in both animals and humans is its
oxidative demethylation, which leads to the formation of t-butyl alcohol
(TBA) and formaldehyde. There is evidence that oxidative demethylation
of MTBE is catalyzed by the cytochrome P450 enzymes (Brady et al.,
1990). TBA was identified in the blood and expired air of rats which
received 14C-MTBE (radiolabeled at the central butyl carbon) by
oral,
intravenous, dermal, or inhalation exposure (Bio-Research Laboratories
Report # 38843; Exxon 1988). Secondary metabolism of TBA results in the
formation of 2-methyl-1,2-propanediol, which is further metabolized to
(alpha-hydroxy isobutyric acid (Bio-Research Reports # 38843-38844, 1990).
Both metabolites were identified in the urine of rats which received
14C-MTBE (Bio-Research Report # 38843-38844, 1990). In vitro
studies
with rat liver microsomes also indicated that TBA may undergo oxidative
demethylation to produce formaldehyde (Cederbaum and Cohen, 1980). A
negligible portion of the administered MTBE dose (1% or less) to the
rats was converted to 14C02 (Table 3).

In another study (API 1984), significantly larger amounts of
14C02 (ca. 7.0% of dose) were reported to have been
expired after
intraperitoneal administration of 232 mg 14C-MTBE/kg
(radiolabeled at
the methyl and the central butyl carbons). Formaldehyde was not
reported in rats which received 14C-MTBE by oral, intravenous,
dermal,
or inhalation exposure (Bio-Research Laboratories Reports # 38842-38845,
1990). However, it was reported that 14C-formic acid accounted
for
most of the MTBE metabolites which were excreted in the urine and feces
(approximately 3 and 1 % of dose, respectively) of rats that received
232 mg 14C-MTBE/kg by intraperitoneal administration (API, 1984).

In vitro metabolism studies revealed that incubating MTBE (5 mM)
with rat hepatic microsomes resulted in the formation of relatively
equimolar amounts of TBA and formaldehyde (Brady et al., 1990).
Additional evidence for formaldehyde formation from MTBE was reported in
studies which investigated the in vitro metabolism of TBA using rat
hepatic microsomes (Cederbaum and Cohen, 1980). Results of these
studies showed that oxidative demethylation of TBA by rat microsomes
resulted in the formation of formaldehyde.

The disposition and toxicokinetics of 14C-MTBE and its major
metabolite, TBA, were compared in mature adult male and female F344 rats
after oral, intravenous, inhalation, and dermal exposure (Bio-Research
Laboratories Reports #38842-45, 1990). 14C-MTBE (radiolabeled at
the central butyl carbon) was administered to groups of rats in a dosing
vehicle of 0.9% saline at a single dose of 40 mg/kg body weight using
gavage, intravenous, or dermal (occluded application) routes. A higher
dose of 400 mg/kg body weight was similarly administered to rats using
oral or dermal application. A summary of the disposition of MTBE is
presented in Table 3.

The major routes of MTBE elimination after oral, intravenous,
intraperitoneal or dermal exposure were via the lungs as expired organic
volatiles and via the kidneys as urinary metabolites (Table 3).
Quantitative analysis of charcoal-trapped expired organic volatiles
showed that the unchanged parent compound accounted for most of the
organic volatiles exhaled by these rats. The portions of MTBE dose
expired as unchanged parent compound and TBA were relatively higher
after oral compared to intravenous exposure (Table 3). Elimination of
MTBE via the lungs was also demonstrated in earlier studies conducted
after MTBE was administered to Fischer rats orally or intravenously
(Exxon, 1988). Additionally, exhalation of unchanged MTBE via the lungs
was reported in mice as the major elimination pathway of MTBE after
intraperitoneal administration at 50, 100, and 500 mg/kg (Yoshikawa et
al., 1994). Pulmonary elimination of parent MTBE in mice ranged from
23% to 69% of the administered dose (Yoshikawa et al., 1994). In a
study where MTBE was administered to CD rats intraperitoneally, a
surprising 91-92% of the dose was eliminated in the expired air as
parent MTBE (Table 3).

TBA exhalation accounted for 3% or less of the administered dose
after all 3 routes of administration (Table 3). Two urinary
metabolites, 2-methyl-1,2-propanediol and (alpha-hydroxy isobutyric acid,
were identified in rats which received MTBE (Bio-Research Laboratories
Reports #38843-38844, 1990). These 2 metabolites are most likely formed
as a result of secondary metabolism of TBA. Fecal elimination of
MTBE-derived radioactivity was minimal and accounted for approximately
1% of the administered dose regardless of the route of administration
(Table 3).

The toxicokinetic parameters of MTBE and TBA depend on the dose and
route of administration (Table 4). While the half-life (t1/2) of
MTBE ranged from 0.45 to 0.62 hour after the low oral and intravenous dose,
it ranged from 0.79 to 0.93 hour after administration of the high doses
(Table 4). On the other hand, the t1/2 of MTBE after
intraperitoneal and dermal application varied from 0.8-1.0 and 0.9-2.3
hours, respectively (Table 4). The t1/2 of inhaled MTBE was
unaffected by dose or repetitive exposure to MTBE and ranged from
0.51-0.63 hour (Table 4). Maximum plasma concentration (Cmax)
was higher among males than females after oral administration of MTBE (Table
4), and the difference was statistically significant only after
intravenous administration. The total plasma clearance (CL) of MTBE was
relatively similar in all animals receiving similar doses of MTBE
regardless of the route of administration and ranged from 273 to 481
ml/hr. TBA, the main metabolite of MTBE, was detected in the expired
air; however, products of TBA's secondary metabolism
(2-methyl-1,2-propanediol and alpha-hydroxy isobutyric acid) were detected
in the urine (Bio-Research Laboratories Report #38843, 1990).

In another study, MTBE metabolism and toxicokinetics were
investigated in F344 rats of both sexes using a 6-hour single exposure
to 400 ppm (calculated doses in males and females are 215 and 293 mg/kg,
respectively) or 8000 ppm (the calculated doses in males and females
were 4220 and 5840 mg/kg, respectively) using nose-only inhalation
(Bio-Research Reports Laboratories #38844-38845, 1990). Using the same
route of exposure, the toxicokinetics of MTBE was also investigated
using repeat 6-hour daily exposure to 400 ppm (calculated doses in males
and females were 245 and 344 mg/kg, respectively) after 14 days of
6-hour daily exposures to unlabeled MTBE (Bio-Research Laboratories
Reports #38844-38845, 1990).

In contrast with routes of elimination when 14C-MTBE was
administered orally, dermally, intravenously, or intraperitoneally, the
main route of post exposure elimination of MTBE-derived radioactivity
after nose-only inhalation was in the urine; the second major route of
elimination was exhalation of organic volatiles (Table 3). However, the
percentage of dose eliminated via each of the 2 routes (lungs and the
kidneys) may actually be different since exhalation of MTBE and TBA
during the 6-hour exposure to MTBE was not accounted for in these
studies. Minimal changes in the disposition of MTBE were observed after
repeat nose-only inhalation exposure to 400 ppm. In both exposure
regimens, approximately 60-70% of the dose was eliminated in the urine
compared to 15-20 % of the dose exhaled in the expired air.
Furthermore, the ratio of MTBE/TBA exhaled in the expired air was lower
in animals which received MTBE via nose-only inhalation compared to
oral, intravenous, or dermal exposure (Table 3). Significant changes in
the disposition of MTBE were observed after exposure to 8000 ppm. A
decline in the portion of the dose eliminated in the urine in
association with an increase in the exhalation of MTBE was observed
(Table 3). In addition, the ratio of MTBE/TBA exhalation increased
after exposure to 8000 ppm as a result of increased exhalation of
unchanged MTBE (Table 3). Further, significant differences in the AUCs
of the 2 doses were observed (Table 4).

In summary, studies of the metabolism, disposition, and
toxicokinetics of MTBE in animals demonstrated that:

MTBE was rapidly absorbed into the circulation from all sites of
administration and distributed to all major rat tissues. Maximum blood
concentrations of MTBE were rapidly achieved (15 minutes after oral
administration vs. 4-6 hours after dermal application). Metabolism and
elimination of MTBE and its metabolites also proceed rapidly regardless
of the route of administration.

Absorption of dermally applied MTBE was somewhat limited and
slower than absorption after other routes. The relative bioavailabilty
of dermally applied MTBE was approximately 20% and 40% at the high and
low doses, respectively. It is likely that dermal absorption of MTBE
may be limited by the high volatility of MTBE.

Elimination of MTBE and its metabolites occurred mainly via the lungs
and the kidneys. Exhaled organic compounds comprised MTBE and TBA and
the ratio of the two components was dose and route dependent. Clearance
of unchanged MTBE via the lungs is likely to be a function of blood/air
partition coefficient and appears to be directly proportional to the
Cmax of MTBE achieved. A small portion of the administered dose was
eliminated as CO2. A negligible portion of the dose was
eliminated in the feces.

Urinary elimination of MTBE-derived radioactivity was also dose- and
route-dependent and two urinary metabolites, 2-methyl-1,2-propanediol
and (alpha-hydroxy isobutyric acid, were identified in the urine of rats
which received MTBE. These 2 metabolites are most likely formed via the
secondary metabolism of TBA. MTBE metabolism to formaldehyde was
demonstrated in vitro using rat liver microsomes (Brady et al., 1990).
However, there is little evidence for the formation of formaldehyde in vivo.
One study reported the elimination of small amounts of formic acid in the
urine and feces of rats that received MTBE by intraperitoneal
administration (API 1984).

From the available experimental evidence, it is proposed that MTBE
undergoes oxidative demethylation via the cytochrome P450 enzymes to
yield TBA and formaldehyde. TBA may be eliminated unchanged in the
expired air or may undergo secondary metabolism resulting in the
formation of 2-methyl-1,2-propanediol and (alpha-hydroxy isobutyric acid;
both are eliminated in the urine. In vitro evidence also suggested that
TBA may undergo oxidative demethylation to produce to formaldehyde.
Identification of CO2 in the expired air of MTBE-treated rats is
indicative that a portion of the administered MTBE dose was subjected to
complete oxidation. It is likely that complete oxidation of MTBE to
CO2 may proceed via the formaldehyde intermediate.

Since MTBE metabolism is catalyzed by the cytochrome P450 enzymes,
concurrent exposure to other environmental chemicals (e.g. gasoline
emission combustion product, cigarette smoke) which are known to affect
this enzyme system may also alter MTBE metabolism and its potential
toxicity or carcinogenicity.

Although human studies showed that exposure to MTBE, similar to animal
studies, leads to the appearance of TBA in blood, no reports of MTBE
metabolism to formaldehyde in humans is currently available. However,
there is no reason to believe that humans would metabolize MTBE in a
qualitatively different manner than animals. In order to assess the
contribution of MTBE metabolites to its overall toxicity, it is
essential to fully investigate the qualitative and quantitative behavior
of MTBE kinetics in both humans and animals. Determination of the rates
of metabolism as well as the internal dose of MTBE, and its main
metabolites (TBA and formaldehyde) will be critical for understanding
the mechanisms of MTBE-induced toxicity. Further, this data will be
essential for extrapolation from animals to humans and for more accurate
assessment of the potential health risks to humans from exposure to MTBE.

Differences in the disposition and toxicokinetics parameters of
MTBE after various doses administered by the same route and differences
after the same dose administered via various routes suggested that there
is a potential for the saturation of MTBE metabolizing enzymes at high
exposure levels and after bolus administration of MTBE.

Studies on the disposition and toxicokinetics of MTBE in
experimental animals after oral, intravenous, dermal, intraperitoneal,
and inhalation exposures suggested that neither MTBE nor any of its
metabolites has the potential for bioaccumulation in animals.

Some evidence of gender differences in the disposition and
toxicokinetics of MTBE is evident from some of the data presented in
Tables 3 and 4. However, the differences were not consistent and
therefore the biological significance of such differences remains to be
established.

TABLE 3: Disposition of MTBE and its metabolites (TBA and
CO2)
in rats after oral, dermal, intravenous, inhalation, and intraperitoneal
exposure.

Route and Dose
of MTBE

Sex

ExhaledMTBEa

ExhaledTBAa

ExhaledCO2b

Urineb

Fecesb

References

Oral

40(mg/kg)

MF

42.050.3

3.03.1

0.140.18

34.727.1

1.190.87

Bio-Res.#38843, 1990

400(mg/kg)

MF

57.162.6

1.41.3

0.200.12

16.010.8

0.280.26

IV

40(mg/kg)

MF

38.042.7

2.52.6

0.120.18

25.724.0

0.830.63

Bio-Res.#38843,
1990

Dermal

40(mg/kg)

MF

4.27.4

0.650.84

0.010.0

6.045.34

0.140.13

Bio-Res.#38843,
1990

400(mg/kg)

MF

1719

0.991.20

0.350.11

11.79.66

0.190.11

Inhalation

400 ppm(6 hr)

MF

10.510.8

5.46.3

0.650.78

62.4962.25

0.760.74

Bio-Res.#38845,
1990

400 ppm(repeat

MF

9.612.8

4.25.9

0.590.61

69.1263.35

0.620.59

800 ppm(6 hr)

MF

20.225.3

5.24.2

1.121.06

40.5833.52

0.751.06

IP

232 (mg/kg)

MF

91.392.0

NANA

8.026.75

2.413.48

0.801.25

API 1984

aExhaled % of MTBE dose at 6 hours after dosing.bUrinary and fecal elimination, as well as CO2
exhalation are presented as % of MTBE dose at 48 hours after dosing.
NA = data not available.

TABLE 4: Toxicokinetic parameters of MTBE and TBA in
rats of both sexes after oral, dermal, intravenous, inhalation,
and intraperitoneal exposure.

E. Biological Measurements in Humans

1. Controlled human exposure studies

To date, three controlled human exposure studies on MTBE have
been conducted. These three studies had similar designs and were aimed
at evaluating internal dose levels resulting from known controlled
exposure and the pharmacokinetics of uptake and decay.

In 1993, EPA and CDC collaborated to examine the internal dose
concentrations of MTBE and its metabolite TBA resulting from exposing
one male and one female volunteer for 1 hour to a 1.39 ppm MTBE
concentration in a controlled environmental chamber (Gerrity et al.,
1993; Prah et al., 1994; Buckley et al., 1995). MTBE and TBA levels in
blood and urine and MTBE in the breath of these two subjects were
determined both during the exposure phase and for 8 hours after the
subjects were removed from the exposure. These subjects had
significantly different body mass - the male weighed 102.5 kg and the
female weighed 66.5 kg.

In both subjects, blood MTBE levels showed a rapid increase during
the exposure phase but did not reach a steady-state plateau after 1
hour. Once the subjects were removed from the chamber, the blood MTBE
levels decreased rapidly indicating that mo st of the internal dose
quickly leaves the body, but in one subject these levels did not return
to pre-exposure levels even after 7 hours. TBA levels in blood also rose
quickly after the exposure began, but there was a much longer plateau
period that extended over the entire 7-hour evaluation. Breath
measurements on these subjects did not return to baseline after 7 hours
in either case (Buckley et al., 1995).

MTBE levels measured in urine were of similar concentration and
followed a similar course to blood MTBE concentrations with a rapid rise
and decay. TBA levels in urine increased to a plateau, which remained
constant for a number of hours, but the rise in these levels appeared to
be delayed for blood TBA. Levels of MTBE in the breath also showed a
similar response to levels in blood and urine with a rapid increase upon
the initiation of exposure. The rate of increase slowed with time, but
did not reach a plateau during this 1-hour exposure period. Thus, these
results show that there is a rapid equilibration among blood, breath,
and urine levels of MTBE. The metabolism of MTBE to TBA is a rapid
process, and even though a large portion of MTBE is exha led unchanged, a
fraction is metabolized to form TBA. Excretion of unmetabolized MTBE
should also be considered as a significant elimination mechanism since
the urine levels of this compound rapidly increase upon exposure.

The MTBE blood level reached after a 1-hour exposure to 1.39 ppm
MTBE in air for the male subject was 8.2 ug/L and for the female
subject, 14.1 ug/L. Thus, the response of MTBE blood levels was 1.7
times higher in the second subject than in the first. The breath levels
in the second subject were also much higher than in the first subject.
After 1 hour, MTBE levels in the breath were 43.0 ug/L in the first
subject and 69.5 ug/L in the second, so that the second subject's
breath MTBE level was 1.6 times higher than the first subject's breath MTBE
level. This figure is in agreement with the blood results. These
limited findings suggest a dependence of the internal dose of MTBE on
body mass that differed in the two subjects by a ratio of 1.5. In
contrast, the level of TBA in blood was similar for these subjects,
reaching maximum levels of 9.5 ug/L and 10.4 ug/L.

Evaluation of the blood and breath MTBE concentrations after removal
from exposure indicates that the decay process over time is
multiexponential. Fitting the blood and breath to a three-exponential
model yields residence times of 10.5 minutes, 75.1 minutes and 31 hours;
and 3.3 minutes, 33 minutes and 8 hours, respectively. Because of the
limited number of samples taken during the decay period, there is a wide
uncertainty in these residence times, but the presence of a
multiexponential decay is clear. This finding explains how MTBE can be
eliminated rapidly from the body immediately after exposure stops and
still not return to baseline 7 hours after the subjects were removed
from exposure to MTBE. Previous studies have discovered this same
finding with other VOCs (Pellizzari et al., 1992). In these previous
studies, the long-term exponential decay was given as evidence of the
deposition of VOCs into deeper body stores, most likely in adipose
tissue. The determination of a long-term exponential decay for MTBE
concentration suggests that, upon exposure, a significant portion of
this compound also deposits in deeper body stores. The studies
performed by the EPA and CDC cited earlier were limited to short-term
exposures to people who had been exposed recently to MTBE. A slowly
eliminated component suggests that, with repeated exposures, the
internal dose of MTBE in the body may remain elevated over a long time
period. There will be short-term spikes in the internal dose of MTBE as
each exposure occurs, but this compound will give a steady-state level
that depends on the frequency of exposure events and the exposure levels
during these events. Pre-exposure levels of MTBE will be indicative of
long-term exposure integration, and samples taken immediately after or
during exposure will show short-term levels. The slow excretion of TBA
suggests that this compound will accumulate even more than MTBE, so that
a substantial internal dose level of TBA may result even when exposure
to MTBE occurs at lower levels or less often.

In 1993, investigators at Yale University (Cain et al., 1993)
measured blood MTBE and TBA levels in four subjects before, during, and
after 1 hour of exposure to 1.7 ppm of MTBE. Measurements were made
for 90 minutes after the end of exposure. This study also showed a
rapid rise of MTBE blood concentrations during the exposure period with
failure to reach equilibrium after 1 hour. MTBE concentrations reached
an average of 17.1 ug/L after 1 hour and decayed after the subjects
were removed from the exposure. The rate of decay appears to be slower than
that reported in the EPA/CDC study in which the half-life was reported
to be approximately 40 minutes; however, after 90 minutes, the MTBE
levels were still a factor of 9 above the pre-exposure level.
Peak MTBE blood concentrations were equivalent to the second EPA/CDC
subject after differences in exposure level are taken into account.

Swedish researchers (Nihlen et al., 1994; Johanson et al., 1995) have
performed a series of controlled chamber MTBE-exposure experiments at
concentrations that were significantly higher than those described in
the first two studies and that lasted for 2 hours. They measured levels
of MTBE and TBA in blood, urine, inhaled air, and exhaled air during and
after exposure of 10 healthy male volunteers to 5, 25, and finally, 50
ppm MTBE vapor during light physical exercise. They monitored blood
levels for 24 hours (48 hours at 50 ppm) after the cessation of
exposure. Low uptake, high post-exposure exhalation, and low blood
clearance indicated the slow metabolism of MTBE.

Blood concentrations of MTBE reached a maximum concentration of
approximately 1.2, 6.5, and 12.5 uM, which are equivalent to 106, 570,
and 1100 ug/L for the 5, 25, and 50 ppm MTBE exposures. These levels
are approximately twice those expected from the previous studies. The
differences can be explained at least in part by the longer exposure
time and by the subjects' light exercise during exposure. The
pharmacokinetics of MTBE in blood was determined to require a
three-exponential fit of the data with half-lives of 10 minutes, 1.5
hours, and 19 hours. These results are in good agreement with the
EPA/CDC results and support the indication that, although most of the
blood concentration of MTBE is quickly eliminated, long-term elevated
levels of MTBE may occur with repeated exposure. The area under the
concentration curves of blood MTBE and blood TBA were linearly related
to the MTBE exposure level, suggesting that the toxicokinetics are
linear up to at least 50 ppm. In contrast to MTBE however, blood TBA
continued to increase during the 2-hour MTBE exposure, then leveled off
and started to decline about 6 hours later. The post-exposure
half-life of TBA in blood was 10 hours.

The study by Nihlen et al. 1994 also found detectable levels of MTBE
and TBA in urine as did the EPA/CDC study. Less than 1% of the absorbed
dose of MTBE was excreted as TBA in urine within 24 hours; this finding
may indicate further metabolism of TBA in humans as has been shown in
rats (Bio-Research Laboratories Report #38845, 1990). Urinary MTBE
levels decreased rapidly after exposure ended, and TBA concentrations
changed during the uptake and elimination phases with a rapid increase,
long plateau phase, and a half-life of 7-9 hours. The presence of MTBE
and TBA in urine and the longer half-life of MTBE in urine compared to
blood suggest that urine may be a useful matrix for determining
internal-dose levels of MTBE and TBA. Since people are more likely to
be willing to give a urine specimen than a blood sample, it is probable
that subject participation would increase. However, since urinary
volume output is variable, a single measurement of TBA or MTBE may not
be as meaningful as a similar measurement in blood. Use of MTBE or TBA
in urine must be fully investigated before this matrix can be accepted
as a sensitive measure of MTBE exposure.

2. Epidemiologic field studies

a. Fairbanks, Alaska

Investigators from CDC's National Center for Environmental Health
(NCEH), in collaboration with the State of Alaska, NIOSH, and EPA,
conducted a two-phase study in Fairbanks, to investigate the
relationship between exposure to MTBE and self-reported health effects
among people who were likely to encounter MTBE either in their
occupation or while commuting to work (Moolenaar et al., 1994). In this
study, 18 workers were examined in December 1992 while the oxygenated
fuels program was in effect (Phase 1), and 28 workers were examined in
February 1993 after the program had been discontinued in late December
(Phase 2). Blood levels of MTBE and TBA were evaluated before subjects
began work and after they completed their work shift. Additional
measurements of blood MTBE and TBA were also made by obtaining blood
samples of office workers (commuters) before and after their commute.

For workers in Phase 1, the median pre-shift concentration of
MTBE in blood was 1.15 ug/L (range 0.1 - 27.8 ug/L), with an
increase to a median post-shift concentration of 1.80 ug/L (range 0.2 -
37.0 ug/L). This difference was statistically significant.
Among workers in Phase 2, median pre-shift MTBE levels were 0.21 ug/L
(range < 0.05 - 4.35 ug/L) and the median postshift concentration
was 0.25 ug/L (range < 0.05 - 1.44 ug/L). The results show
that blood MTBE levels increased during the work-shift while MTBE was
being used as a fuel oxygenate, and blood levels were lower after the
cessation of the oxygenated fuels program.

In Phase 1, the blood MTBE levels among commuters changed from 0.18
ug/L before they commuted to 0.83 ug/L afterward. In Phase 2, the
blood MTBE levels before the commute (0.09 ug/L) and after the commute
(0.10 ug/L) were not significantly different (t-test). Since many
people commute to work each day, the increase in MTBE internal dose levels
among commuters suggests that a large proportion of the public in
locations where MTBE is present in gasoline receive a measurable
exposure to this compound.

It is significant that blood MTBE levels were not below detection
limits for either workers or commuters before the work shift or morning
commute (the period of MTBE exposure). The median blood MTBE level in
15 participants in the Third National Health and Nutrition Examination
Survey was below the detection limit of 0.05 ug/L (Ashley, 1993).
Elevated levels measured before exposure support the results of
controlled human exposure studies because they suggest that MTBE levels
in the body have not returned to baseline even after 16 hours. Thus,
daily exposures may yield elevated levels for a time much longer than
the estimated half-life.

In the Fairbanks study, there was a significant correlation between the
difference between pre- and post-shift blood MTBE levels among workers
and in workplace air concentrations. For these workers, the median air
concentration of MTBE was 0.1 ppm and blood levels showed a
statistically significant mean difference between pre- and post-exposure
samples of 0.65 ug/L. Among some of the workers whose exposure was
greater, air levels of 0.55 ug/L yielded a difference between pre- and
post-exposure samples of 10 ug/L. This finding is in agreement with
the EPA/CDC controlled-chamber exposure study in which air levels of 1.4 ppm
MTBE in the chamber yielded 11 ug/L MTBE in the blood. For the
workers, the internal dose levels of MTBE were higher, since the workers
were exposed for 8 hours rather than 1 hour as in the controlled exposure
studies. Blood levels do not increase linearly as time increases,
because the internal dose levels begin to plateau after about 1 hour.
Thus, the internal dose levels of MTBE are a function of both exposure
time and exposure concentration and not just their product as exposure
is normally defined.

b. Stamford, Connecticut

In 1993, CDC, in collaboration with health officials from the state
of Connecticut, examined MTBE blood levels among 14 commuters and 30
other people who worked around traffic or automobiles (White et al.,
1995). The highest blood MTBE levels were found among gas station
attendants (median = 15 ug/L, range = 7.6 to 28.9 ug/L). People
who worked in car repair shops had highly variable blood MTBE results
(median = 1.73 ug/L, range = 0.17 to 36.7 ug/L), and commuters had
the lowest levels of MTBE in blood (median = 0.11 ug/L, range =
<0.05 to 2.60 ug/L). Blood levels of MTBE were highly correlated
with PBZ samples of MTBE.

The measurements of blood MTBE levels among car-repair workers in
Stamford, Connecticut, is in excellent agreement with the levels found
among the workers in Fairbanks. Thus, this segment of the population
receives a substantial MTBE exposure even in a region of the country
that does not have Alaska's extreme conditions. Internal dose levels of
MTBE among commuters were lower than those found in Fairbanks. This
difference may be due to the differences in driving habits during the
periods when these studies occurred. The first phase of the Fairbanks
investigation took place during the winter when commuters are more
likely to not open their car windows, thus reducing the free flow of
fresh air into the vehicle. In the Stamford study, samples were
collected at the beginning of spring, when ventilation through the
vehicle may have reduced the extent of exposure. Since the oxygenated
fuels program is chiefly in effect during the winter months, the weather
may have had a significant effect on the extent of exposure of commuters.

3. Summary: Biological exposures in humans

To date, only a limited number of studies of biological levels
of MTBE and TBA have been performed in either controlled human exposure
or epidemiologic field studies. The measurement of concentrations in
blood, breath, and urine indicates that sampling from these matrices
gives equivalent results and can be used to evaluate exposure. The
relationship between human inhalation exposure and blood level appeared
proportional to the relationship observed in rats.

Studies in humans of MTBE metabolism suggest that, with repeated
exposures, levels of both MTBE and TBA may be elevated in biological
fluids for a time much longer than the half-life with TBA doing so to a
larger extent as a result of the longer residence times for this
metabolite. Exposure will be integrated over a period of days for MTBE
and possibly longer for TBA when repeat exposure to MTBE occurs. Thus,
there will be repeated spikes in the internal dose immediately after
exposure followed by a steady-state level between exposure events.

By carefully timing sample collection with exposure events, it is
possible to evaluate exposure and to gain insight into the duration of
either short-term or long-term MTBE levels. Internal-dose
concentrations found among workers and commuters indicate that both
highly exposed populations and the commuting public experienced
increases in internal-dose levels of MTBE when the oxygenated fuels
program was in effect.

A. Epidemiologic Studies

Before considering specific epidemiologic investigations, it is
useful to consider the context in which they were conducted and the
questions that anecdotal reports of health complaints have generated.
In fact, much of the epidemiologic work that has been done on oxygenated
gasoline was stimulated by anecdotal reports. Many of the surveys that
have been conducted on health symptoms, however, were designed as
screening or exploratory studies and were not designed to test specific
hypotheses. These surveys, however, have helped clarify what those
hypotheses should be.

1. Anecdotal reports

a. Reports of symptoms among motorists:

In November 1992, public concern about oxygenated fuels in
Fairbanks, Alaska, arose shortly after the beginning of the oxygenated
fuel program. People's awareness of the oxygenated fuel program could
have been increased not only from intense media coverage, but also from
a lowered threshold for perception of an unpleasant odor of the gasoline
in Alaska (Smith and Duffy, 1995). Call-back interviews to 34 people
who had reported complaints to a media-publicized citizen hotline
reported these symptoms more frequently: primarily headache, cough,
nausea, dizziness, burning sensation of the nose or throat, irritation
of the eyes, and a sensation of spaciness or disorientation (Beller and
Middaugh, 1992). Subsequent investigations that winter by the Alaska
Department of Health and by others the following spring focused
primarily, but not exclusively, on these seven symptoms.

In response to considerable media coverage and public concern about
oxygenated fuel containing MTBE in Missoula, Montana, during the
1992-1993 winter season, the Missoula City-County Health Department
conducted a survey of local physicians in January 1993 to screen for
possible health effects (Missoula City-County Health Department, 1993).
A total of 18 physicians' offices were contacted by telephone. Twelve
of the offices had received complaints from patients specifically
regarding oxygenated fuels; the symptoms reported included many of the
same acute effects that had been reported in Alaska, as well as others,
including an exacerbation of symptoms among people with asthma. The
next winter, the oxygenate used was ethanol rather than MTBE, and
public concern over this issue essentially disappeared (E. Leahy,
Missoula City-County Health Department, personal communication, 1995).

Several other state health departments have also received some
health complaints related to oxygenated gasoline from citizens similar
to those received in Alaska. Passive reliance on reports of citizen
complaints to health departments, however, is a highly insensitive and
potentially misleading measure of health effects related to gasoline and
its components. State health departments do not routinely conduct
surveillance for complaints related to gasoline or other environmental
exposures. In addition, citizens would not necessarily report health
complaints of a nonspecific and noninfectious nature, such as a
headache, to the health department, even if they believed symptoms were
caused by gasoline or some other environmental agent. If citizens
did report such symptoms to other organizations, such as to a poison
control center, local medical providers, or the local environmental
control agency, these reports easily might not be communicated to health
authorities.

For example, in the winter of 1992-1993, the Colorado Department of
Public Health and Environment received three complaints about the health
effects of, or odors related to, oxygenated gasoline (Livo, 1995). That
same winter, a private citizen in Colorado Springs actively collected
reports from more than 50 people about health complaints as a result of
exposure to MTBE (Pat McCord, personal communication, 1993). In the
absence of concerted efforts to identify and verify health complaints in
the community, analyses of the number of complaints received by
different health departments or trends in the number of complaints
received over time are difficult to interpret.

A recurrent hypothesis put forth by many of the people who have
expressed concerns about the health effects of MTBE, and one that EPA
acknowledges (USEPA, 1993; 1994), is that some people may be more
sensitive than others to MTBE. To explore this possibility further,
investigators in New Jersey interviewed 14 people with multiple chemical
sensitivities (MCS) and 5 people with chronic fatigue syndrome (CFS)
(Fiedler et al., 1994). The investigators also interviewed six
otherwise healthy people as control subjects, but comparisons between
groups composed of so few people are difficult to interpret.
Qualitatively, people with MCS and CFS reported symptoms previously
associated with MTBE with apparently greater frequency than they
reported other symptoms and more frequently than control subjects, but
these symptoms were reported to develop in locations such as shopping
malls where exposures to MTBE were presumed to be fairly low.

b. Reports of symptoms among workers exposed to oxygenated gasoline

API collected information from 16 member companies regarding
complaints that their workers had reported to them (Raabe, 1993). The
principal complaints were headache, dizziness, and nausea. From 1984 to
1991, the number of complaints received each year ranged from 0 to 5,
but jumped to 45 in 1992. In 1992, most of the worker complaints were
made in October and November, at the start of the oxygenated fuel
season. Although the production of MTBE had been increasing before 1992
and increased only about 12 percent from the previous year, exposures of
some workers in specific occupational categories might have changed at
the beginning of the oxygenated fuel season. Given that these 16
companies made more than 50% of the gasoline sales in the country
and that gasoline itself can lead to a variety of health complaints, the
few complaints recorded each year suggest that this reporting system is
not sensitive.

Refinery workers who are members of the Oil, Chemical, and Atomic
Workers Union and exposed to MTBE at several refineries have reported
many of the same symptoms, as well as others not previously associated
with MTBE or gasoline, that were reported among motorists in Fairbanks
when MTBE was first introduced as an oxygenate (Medlin, 1995). The most
common complaints were headaches, sinus problems, fatigue, and shortness
of breath.

2. Epidemiologic investigations

a. People occupationally exposed to oxygenated gasoline:

As part of the follow-up to the initial concerns about oxygenated
gasoline in Fairbanks, CDC collaborated with state health officials on
an investigation of people who had different opportunities to be
exposed to oxygenated gasoline (Moolenaar et al., 1994). These
investigations included fairly small numbers of adults, including
mechanics, who worked around traffic and motor vehicles while the
oxygenated fuel program was in place and again after it had been
suspended. This investigation included obtaining both blood and air
measurements of MTBE and other gasoline components, and these
measurements are discussed in the section of this assessment on exposures.

The workers also completed questionnaires that asked about the
occurrence of 15 different health symptoms, the 7 that had been
frequently reported by citizens and 8 others. Questions focused on
symptoms that had occurred for the first time or with greater frequency
since the beginning of the oxygenated season and which the respondent
did not attribute to having a cold or the flu. The questionnaires were
administered in person to the workers at the end of the work shift. By
necessity, the workers studied were few and were not randomly selected,
so the actual prevalence figures for specific symptoms may not reflect
the prevalence of these symptoms among other workers who were similarly
exposed. Certain symptoms, especially headache, eye irritation,
and a burning sensation of the nose and throat, were much more common
when the oxygenated program was in operation than after the program had
been suspended. This investigation, however, could not determine
whether exposure to MTBE in oxygenated gasoline
actually was the cause of increased symptom reporting.

After the investigations in Fairbanks and with the support of the EPA,
CDC undertook two subsequent investigations in collaboration with state
health officials in Stamford, Connecticut (CDC, 1993b), and in Albany,
New York (CDC, 1993c). The follow-up investigations were intended to
duplicate the methods used in Fairbanks in a city that had used
oxygenated gasoline but had not experienced media publicity and had not
received complaints from local citizens (Stamford), and in a city that
had not participated in the oxygenated fuel program (Albany). One
important distinction between the investigations in Fairbanks and these
two subsequent investigations is that they occurred at different times
of the year under different climatic conditions.

As in the Fairbanks investigations, the investigations in
Stamford and Albany included workers, such as auto mechanics and
gasoline service station workers, who had opportunities for exposure to
gasoline and motor vehicle emissions. The questionnaires were designed
to be nearly identical to those used in Fairbanks and also focused on
symptoms that the respondent did not attribute to a cold or the flu.
Air and blood measurements of MTBE and other gasoline components from
these two investigations are discussed in the exposure-profile section.

Because of time and resource constraints, both the Stamford and
Albany investigations relied on convenience samples of relatively small
numbers of workers in different job categories. In both locations, the
occupationally exposed workers were predominantly men. Despite these
and other limitations, these investigations provided good comparative
data on exposures and doses and gave a qualitative indication of the
magnitude of symptoms experienced by some workers in these two cities.

The small numbers of workers and the nonrandom nature of their
selection, however, limit the interpretation of specific prevalence
figures. Qualitatively, the prevalence of the most common symptoms,
such as headache and cough, occurring over the last month, were not
appreciably higher among men who worked around cars and gasoline in
Stamford than men with similar occupations in Albany, where exposure to
MTBE was generally much lower. In addition, the prevalence of these
symptoms was lower among workers in Stamford than among workers in
similar job categories in Fairbanks during the oxygenated fuel
season.

In Stamford, job category alone was not a good indicator of exposure;
both blood and air measurements revealed that people who worked as
mechanics and at gasoline stations had highly variable exposures to MTBE
(White et al., 1995). Blood measurements were available for only 30
workers; of these workers, the 8 with the highest blood levels of MTBE
were significantly more likely to report 1 or more of the 7 key symptoms
on that day than were the remaining 22 workers (OR = 21.0, 95% CI = 1.8
- 540). These workers with the highest exposures to MTBE were also more
heavily exposed to gasoline. There was no comparable reference group
of workers in Stamford exposed to gasoline not containing MTBE.

In April and May 1993, researchers at EOHSI interviewed garage
workers in two parts of New Jersey: the northern part of the state near
New York City where oxygenated fuel was still in use and where exposures
to MTBE were higher, and the southern part, near Philadelphia, where use
of oxygenated fuels had been discontinued at the end of February (Mohr
et al., 1994). These garage workers were all employed by the State of
New Jersey, had similar backgrounds and training, and worked under
similar conditions. Workers were asked about the prevalence of symptoms
during the last 30 days and at the beginning of the work shift and again
at the end of the work shift. The questionnaire asked about the same
symptoms that had been included in other investigations, but the wording
was different and did not ask the respondent to identify symptoms not
attributed to colds.

Overall, this investigation did not report major differences in
symptom reporting between these two groups of workers in New Jersey with
presumably different levels of exposure to MTBE. Comparisons between
people who pumped gasoline in the north and south, matched by age, also
did not demonstrate a difference in the prevalence of symptoms, but this
analysis was made on the basis of information from only 11 workers in
each area. The authors noted that this investigation was conducted at
the end of the oxygenated fuel season; any change in the occurrence of
symptoms, if caused by exposure to oxygenated fuels, would probably have
been more noticeable at the beginning of the oxygenated fuel season.

b. Motorists exposed to oxygenated gasoline

In December 1992, the Alaska Department of Health and Social
Services conducted surveys of taxi drivers, health care workers, and
university students at different locations in Fairbanks and Anchorage,
Alaska (Beller and Middaugh, 1992; Chandler and Middaugh, 1992). The
total number of people interviewed was fairly modest: 162 in Fairbanks
and 203 in Anchorage. In both cities, about a quarter of the people
interviewed met the case definition (an increase in headaches or two or
more other specific complaints), and this proportion was slightly higher
for taxi drivers and slightly lower for university students. This
survey documented that people in both cities were self reporting an
increase in certain symptoms after the introduction of oxygenated
fuels but provided little information to judge the true prevalence of
such symptoms in the broader population or whether such symptoms were
actually caused by oxygenated gasoline.

Two telephone surveys also were conducted in Fairbanks during
the oxygenated fuel season (December 1992) and after the program had
been suspended in February 1993 (CDC, 1993a). Only 41 people were
interviewed in December; the most prevalent symptoms they reported were
eye irritation, headache, and a burning sensation of the nose or throat,
each reported by a third of the people interviewed. In February, 100
people were interviewed, and the prevalence of headaches and other
symptoms was much lower. Given the quasi-random sampling and small
number of people interviewed, the estimate of the prevalence of
different symptoms is fairly crude in both surveys.

In Anchorage, the medical officer examined records for
outpatient visits among state employees, retirees, and dependents in
Anchorage and Fairbanks for the winter of 1992-1993, when oxygenated
gasoline was in use, compared with such visits the previous two winters
(Gordian et al., 1995). The report of this analysis is brief and
difficult to evaluate fully. The authors report that there had been an
epidemic of headaches in January 1992 due to a viral illness.
Therefore, a comparison in visits for headaches between the winter of
1992-1993, when oxygenated fuels were first in use, and the previous
winter would have been unlikely to measure an increase in headaches.
Compared with the winter of 1990-1991, however, the winter of 1992-1993
showed both Anchorage and Fairbanks reportedly experiencing a 40%-50%
increase in visits for headaches, but the confidence interval for this
estimate was fairly wide and certainly does not rule out chance as an
explanation for this increase. As the authors point out, however, the
data that were used for this analysis were limited in their ability to
measure the sort of transient health symptoms that had been reported as
related to MTBE exposure.

Nonoccupationally exposed motorists had also been included among
the people interviewed in both the Stamford and Albany investigations
(CDC, 1993b; CDC, 1993c). Since both investigations relied on
convenience samples rather than on random samples of motorists, the
people interviewed may not have been representative of the larger
population in either city. In particular, unemployed and retired people
were not included among the motorists interviewed. In both cities, the
most common symptoms were headaches and cough, and the reported
prevalence of these and other symptoms was fairly similar among
motorists in both cities.

During the winter of 1994-1995, the Alaska Department of Health
conducted a weekly random telephone survey of 100 adult residents in
Anchorage, for 16 consecutive weeks, to identify possible health
problems related to the use of ethanol as an oxygenate in fuels (Egeland
and Ingle, 1995). Responses were grouped by time period to correspond
to periods when ethanol containing gasoline was not in use, was in use,
or was being phased in or phased out. The survey questionnaire asked
about the presence of symptoms over the previous week, and persons who
reported illnesses over the past week that could have accounted for
symptoms (fever, sweats or chills, or cold or flu) were excluded. The
results indicated a much lower prevalence of symptoms than had been
reported 2 years earlier when MTBE was used in gasoline, and the
reported prevalence of symptoms remained fairly similar across the
different time periods.

c. Motorists exposed to reformulated gasoline

In response to public concern over reformulated gasoline, the
Department of Health in Wisconsin conducted, during the winter of
1994-1995, a random telephone survey of symptoms among residents in
Milwaukee, Chicago (which also used reformulated fuels but had not
experienced the same intense media coverage of the issue), and areas in
Wisconsin that did not use reformulated fuels (Anderson et al., 1995a).
Although the focus of this study was reformulated fuels, not oxygenated
fuels, it is relevant since the oxygenates used were MTBE, ETBE, and
ethanol, and both MTBE and ethanol are also used in oxygenated fuels,
although at higher concentrations than in reformulated gasoline.

Approximately 500 people in each location were interviewed in
February and March 1995 by using a questionnaire that was based in part
on the questionnaires that had previously been used in Alaska. In
particular, the questionnaire asked whether the respondent had
experienced any unusual health symptoms unrelated to a cold or the
flu. Overall, the survey found that people in Milwaukee reported a
higher prevalence of unusual symptoms than did residents of other areas
of the state or of Chicago. Several other findings are of relevance in
evaluating this increase. First, every symptom was elevated in
Milwaukee, not just symptoms that had previously been associated with
gasoline or chemical solvents. Second, the symptom prevalences were not
elevated in Chicago compared with such prevalences in areas of Wisconsin
where reformulated gasoline was not used, although Chicago was also
using reformulated gasoline. Third, although people were asked to
report unusual symptoms, the symptoms reported in Milwaukee were more
likely to be associated with having had a cold or the flu, smoking
cigarettes, or being aware of reformulated gasoline than were symptoms
reported by people in Chicago or the rest of Wisconsin.

Although well conducted, the study has many limitations. It was a
cross-sectional survey of symptoms but did not measure exposure.
Participation rates were not high, particularly in Chicago. The number
of people interviewed, although substantially greater than in other
surveys of health effects and oxygenated gasoline, still did not provide
sufficient statistical power to measure modest elevations in symptoms or
to look carefully at specific population subgroups.

The study underwent peer review by the Environmental Committee of the
Association of State and Territorial Health Officers (ASTHO) and other
scientists not associated with the study (Anderson et al., 1995a). This
committee concluded that the study was well conducted, had limitations,
and that it had not ruled out the possibility that some people might
have a greater sensitivity to reformulated gasoline. All but one of 16
committee members agreed with the statement that the study did "not
support a conclusion that reformulated gasoline is associated with
widespread or serious adverse health effects."

The same telephone survey questionnaire was subsequently
administered to 1339 persons who had called Wisconsin government
agencies with health concerns regarding reformulated gasoline (health
contacts) between January 26 and March 17, 1995, had provided names and
telephone numbers, and agreed to complete the questionnaire when later
contacted again (Anderson et al., 1995b). Compared to residents of
Milwaukee who had participated in the random telephone survey, the
health contacts were slightly older, more likely to be white, male, own
a car, and commute more than one hour per day, and they were twice as
likely to be retired. In addition, the health contacts were less likely
to smoke cigarettes and were more likely to have been diagnosed as having
allergies, although they were not more likely to have been diagnosed as
having asthma. None of these factors had been observed to be
significant predictors of unusual health complaints in the random
telephone survey. Health contacts also were more likely to have seen
various news stories about MTBE than other Milwaukee residents.

3. Summary of epidemiologic studies

All of these studies were highly responsive to public concerns
and creative in coping with serious limitations in time and resources.
The same limitations apply to many of the studies described: inadequate
sample size, potential bias in sample selection, inadequate exposure
information, and reliance on highly subjective measures of effect. The
few epidemiologic studies have focused mostly on men employed as
mechanics or gasoline service-station attendants who already were
exposed to higher levels of gasoline and other chemicals than those
experienced by the general public. However, these exposures were lower
than exposures which occur among workers in some other occupations, such
as gasoline refineries or transportation. In Fairbanks, Alaska,
complaints of health symptoms were fairly common among a nonrandom and
small sample of workers while oxygenated gasoline containing MTBE was in
use, but these complaints essentially disappeared after the oxygenated
gasoline program was suspended. In two small-scale field studies
conducted in the spring in Stamford, Connecticut, and in New Jersey,
most workers reported no adverse health effects related to oxygenated
gasoline. Some workers did report health complaints, but some evidence
suggested that the prevalence of such complaints was similar for
exposure to oxygenated gasoline compared with conventional gasoline.
Community-based surveys in Fairbanks, Alaska, and Milwaukee, Wisconsin,
indicated that while most people reported no increase in acute
health symptoms after exposure to oxygenates in gasoline, a substantial
number of people attributed headaches and other health complaints to the
presence of oxygenates in gasoline. Evidence from Milwaukee suggests
that increased media attention to this issue could have been one factor
in the greater likelihood of citizen reports of health complaints in a
few communities (e.g. Missoula, Montana; New Jersey) compared with the
likelihood of reports of such complaints other communities.

Taken together, these studies suggest that most people do not
experience adverse health effects from MTBE in gasoline, but the studies
cannot rule out the possibility that some people do experience more
acute symptoms from exposure to oxygenated gasoline than to conventional
gasoline. Many basic questions, such as the relative importance of
individual characteristics, exposure situations, and factors other than
oxygenates for the occurrence of various health symptoms remain. Thus,
a causal association between acute health effects and exposure to MTBE
or other oxygenates in gasoline in a relatively smaller proportion of
persons has not been demonstrated but cannot be ruled out on the basis
of the limited epidemiologic studies that have been conducted to date.
More definitive epidemiologic studies are needed to assess fully the
connection between exposure to oxygenates in gasoline and acute health
effects.

4. Questions not yet addressed by epidemiologic studies

Some of the questions that could be addressed by further
epidemiologic research include the following:

Does exposure to MTBE in gasoline increase the occurrence of acute
symptoms or other acute effects compared with exposure to gasoline that
does not contain MTBE?

Who is experiencing acute symptoms related to oxygenates in
gasoline, and what are the most important risk factors for these
symptoms? What is the role of public awareness about oxygenated fuels?

Can an exposure-response relationship be defined (i.e.,
between symptom reporting and personal exposure to MTBE) in "sensitive"
individuals?

Does a person's health status affect the perception of acute
symptoms from gasoline containing MTBE or other oxygenates? What role
does odor perception play in the occurrence of symptoms?

When do symptoms occur (during refueling or during
commuting?), and is there a consistent pattern? Do some people develop
sensitivity or tolerance over time?

What is the prevalence of symptoms among people who have higher
exposures to gasoline containing MTBE or other oxygenates, and how does
that prevalence compare with the prevalence of symptoms among people
exposed to lower levels of exposure or among people exposed to gasoline
only?

B. Human Experimental Studies

To help address the question of a direct causal relation between
MTBE exposure and health symptoms, studies of the sensory, symptomatic,
cellular, and eye responses of healthy human subjects exposed to MTBE in
air in a controlled exposure chamber have been conducted.

In EPA's investigation (Prah et al., 1994), 37 healthy, nonsmoking
subjects (18 women, 19 men) from 18 through 35 years of age were
studied. Each subject was exposed once for 1 hour to clean air and to
1.4 ppm "pure" MTBE in air on different days. The temperature and
relative humidity in the chamber were maintained at 24°C and 40%,
respectively. The endpoints selected for the EPA study were based on
the observation that the symptomatic reports from Alaska resembled the
types of symptoms associated with low-level organic solvent exposure.
The endpoints for the EPA MTBE study can be divided into four categories:

Densitometric indicator of eye redness (measured before
and immediately after exposure)

Noninvasive tear-film breakup (measured before and immediately after
exposure)

Impression cytology (before and 20 hours after exposure)

Types and numbers of epithelial and inflammatory cells

Biochemical mediators of inflammation

Before exposure testing, each subject had a determination of his or
her individual odor threshold for MTBE in water. Of the subjects in the
study, 76% correctly detected the presence of the odor of MTBE in water
at a concentration of 0.18 ppm. These data compare reasonably well with
data from another study reporting a detection threshold of 0.13 ppm
(Clark, 1993). Thus, it can be assumed that most of the subjects
studied had normal odor thresholds for MTBE.

There was no significant effect of MTBE on the overall
qualitative rating of air quality or on reporting of headache and nasal
irritation symptoms using either the computerized questionnaire or the
EOHSI questionnaire. The neurobehavioral test battery also showed no
effect of MTBE exposure. None of the markers of nasal and eye
inflammation showed a statistically significant difference in the
response to MTBE exposure from that of clean air. MTBE exposure also had
no statistically significant effect on eye redness or on tear-film
breakup times. The primary hypothesis tested in this protocol was that
MTBE would cause changes in the reporting of symptoms of headache, nasal
irritation, air-quality perception, and odor strength perception. Power
calculations performed on the symptom data showed that (at p<0.05),
there was adequate power to detect a 0.5 point change (on a five-point
scale) with 90% power for odor-detection level, headache, and nasal
irritation. A 0.25 point change in nasal irritation could have been
detected with 80% power. Thus, the study had adequate statistical power
to detect MTBE-related changes in symptoms if they had been present.

Investigators at Yale University (Cain et al., 1995) replicated the EPA
study (Prah et al., 1994). A total of 43 subjects (22 men, 21 women)
from the ages of 18 through 34 years participated. All of the endpoints
studied by the EPA investigators were also studied by the Yale
investigators, although slightly different methods for measuring eye
redness, tear-film breakup times, and eye inflammation were used. The
MTBE exposure concentration was slightly higher in the Yale study (1.7
ppm). In addition to single exposures to clean air and to MTBE for 1
hour at 75°F, each subject in the Yale study also underwent a 1-hour
exposure to a complex mixture of 16 VOCs commonly found in gasoline (C4,
C5, and C6 saturates; and C4 and C5 olefins). This exposure to a
surrogate gasoline served as a positive control for the MTBE exposure.
In the pilot phase of investigation, the surrogate gasoline was found to
have no detectable odor. Consequently, isopropyl mercaptan (the odorant
used in natural gas) was added to provide an unpleasant odor. The total
VOC concentration of the atmosphere was 7.1 ppm. Besides the addition
of the VOC exposure and the slightly higher MTBE exposure concentration,
the only other major difference between the EPA and Yale studies was
that, in the latter study, exposures of a given individual were
separated by only 3 days, as opposed to 1 week in the EPA study (in
which each person received both an air and an MTBE exposure).

When MTBE exposure was compared with clean air exposure, Cain et
al. (1995) found essentially the same results as found in the EPA study
(i.e., MTBE exposure had no statistically significant effects on
symptoms, the neurobehavioral test battery, nasal inflammation, eye
inflammation, eye redness, or tear-film breakup times). Results of
statistical power calculations on the Yale symptom data by EPA (House,
1993) were similar to those of the EPA study. Likewise, for the
objective measures, means of the MTBE and control groups were similar
and standard deviations were small. Thus, the design of the study was
sufficiently robust to have confidence in the negative outcome. When
the VOC exposure was compared with clean air exposure, it was found
that the VOC exposure caused an increase in inflammatory cells in the
nasal lavage on the day following exposure (Cain et al., 1995). This
finding was consistent with previous work done at the EPA laboratories
(Koren et al., 1992).

More recently, a human experimental study was conducted at the
Swedish National Institute of Occupational Health (Nihlen et al., 1994;
Johanson et al., 1995), in which 10 healthy male volunteers, ages 23-51
years old, were exposed during light exercise for 2-hour periods to
three successive concentrations of MTBE -- 5, 25, and 50 ppm. Researchers
measured eye-blinking frequency, conjunctival epithelial damage, eye
redness, tear-film breakup time, acoustic rhinometry, nasal/mouth peak
expiratory flow, and inflammation markers in nasal lavage, along with
subjective ratings of "discomfort, irritative symptoms, and CNS
effects." Other than a slight, marginally significant increase in the
indicator "nasal swelling," which was not concentration-related, no
significant irritant or subjective effects of MTBE exposure were found.
However, subjective ratings indicating detection of a solvent odor were
highly significant.

Summary

Taken together, these three studies of MTBE exposure among human
volunteers provide a consistent picture. They show that controlled
exposure to pure MTBE in air under laboratory conditions (around 24oC)
did not cause increased symptoms or any notable adverse effects (e.g.
irritation, behavioral changes) among healthy adult subjects. These
findings do not rule out the possibility that a subpopulation of people
in the general population may be especially sensitive to MTBE alone or
in gasoline or that effects might be associated with exposure to
evaporative or combustion emissions from oxygenated gasoline or with
some other variable (e.g., cold temperature, odor, concurrent illness).

Further experimental studies are needed to determine whether
measurable symptoms can be induced in human volunteers with
self-described sensitivity to oxyfuels. If such effects can in fact be
induced through exposures to gasoline-oxygenate (e.g., MTBE) mixtures
under controlled conditions, then it may be possible to identify the
variables contributing to the occurrence of such reactions. If
indicated, these studies might be extended to investigate the effects of
combustive as well as evaporative oxyfuel emissions, although ethical
and other considerations may make it difficult or impossible to
administer any but relatively low exposure levels of such mixtures.
Experimental studies could also be devoted to quantifying the
concentration-response relationship for any effects. Such studies could
also examine variables such as odor, temperature, chemical sensitivity,
smoking, alcohol use, and nonphysiological factors that might contribute
to the perception of acute symptoms in some individuals.

C. Neurotoxic Potential of Methyl Teritary Butyl Ether

Subjective complaints from the human population include general
items such as headaches, dizziness, and nausea. Although experimental
animals do not permit assessment of such subjective complaints, they do
allow for an evaluation of the neurobehavioral functioning of the
organism. While MTBE displays neuroactive properties at sufficiently
high doses, studies in animals have not provided evidence of overt
neurotoxicity due to MTBE exposure at air concentrations from 100 ppm to
3000 ppm MTBE (Burbacher, 1993; Costantini, 1993). The characteristic
nervous system response to MTBE at high levels is central nervous system
(CNS) sedation with a domination of anaesthetic effects seen at very
high levels of exposure. The sedative effects are transient and usually
dissipate within a couple of hours following cessation of exposure.
Effects on cognitive functioning have not been examined in animals
exposed to MTBE.

1. Acute Exposure

The acute inhalation concentration of MTBE needed to produce death in
50% of rats (LC50) has been calculated to be between 18,000 ppm and
39,500 ppm. The oral dose needed to produce death in 50% of rats (LD50)
is approximately 3.9 g/kg with lethal exposure accompanied by ocular and
mucous membrane irritation, ataxia and signs of CNS depression (ARCO
Chemical Company [ARCO], 1980). Inhalation exposure to MTBE (30,000
mg/m3, 8300 ppm) produced signs of respiratory sensory irritation
(altered frequency and pattern of tidal breathing) in mice (Tepper et
al., 1994). These effects were transient and dissipated after
exposure. Studies conducted in non-human primates suggest that exposure
to 3400 and 4800 ppm MTBE (6 hours/day) produces no toxic signs, whereas
exposure to 8500 ppm produces ataxia and signs of CNS sedation within 3
days (Industrial Bio-Test Lab, 1970).

Evaluation of acute neurotoxic effects has been conducted in rats
following a single 6-hour exposure by inhalation to 0, 800, 4000, and
8000 ppm MTBE (Gill, 1989). Evaluation of neurotoxic effects was
conducted within 1 hour after exposure and at 6 and 24 hours to examine
the transient nature of the effect. According to clinical observation
methods, there was no evidence of toxicity after exposure. When
behavioral function was evaluated using a systematic screening battery
of tests known as the Functional Observational Battery (FOB), distinct
signs of toxicity were evident. Within 1 hour, a concentration-related
increase in the incidence or severity of ataxia and gait change was
observed in the groups exposed to 4000 ppm and 8000 ppm MTBE. Exposure
to 8000 ppm produced labored respiration, lacrimation, decreased muscle
tone, decreased rectal temperature, decreased treadmill performance, and
an increased hind limb splay. In a few of the animals that received the
high dose, there was evidence of decreased pupil size, loss of pupil
response to light, decreased startle response, and decreased reflex
response to toe pinch. All biologically relevant findings were
restricted to the 1-hour post-exposure test session with no effects
continuing to the 6-hour evaluation period. Automated measurements of
motor activity were conducted during the hours immediately after
exposure. Activity measurements were of high counts and large
variability suggesting that the specific apparatus used had a photocell
placement and sensitivity that measured movements other than just
ambulatory. In male rats exposed to 8000 ppm MTBE, a decrease in mean
activity level was evident within the initial 10 minutes of the test
session, followed by a transient increase and a subsequent decrease
during the 5-hour test session. For animals receiving the
two lower doses (800 and 4000 ppm), mean activity was elevated relative
to control levels during the initial 10 minutes, reflecting either a
low-dose stimulant effect or an exaggerated recovery from anesthetic
effects similar to what is seen with the biphasic response to other
substances with sedative properties. Initially, females showed a
similar pattern of depressed activity. However, within 1 hour, activity
had returned to control levels. These changes in behavioral
functioning, detected by both the FOB and motor activity, are consistent
with transient CNS sedation.

2. Subchronic Exposure

In the initial dose-ranging study, examination of the effects of
subchronic exposure, 6 hours a day for 13 consecutive days, to MTBE at
concentrations of 0, 2000, 4000, and 8000 ppm revealed clinical signs of
hypoactivity, ataxia, and perioccular irritation for rats and mice in
all groups during exposure (Dodd and Kintigh, 1989). Immediately after
exposure to 8000 ppm MTBE, the rats displayed clinical signs of toxicity
characterized in the FOB as decreased startle and pain reflexes and
muscle tone.

After the 13-day dose-ranging study was completed, Dodd and
Kintigh (1989) evaluated neurotoxicity after 13 weeks of repeated
exposure to MTBE vapors. Male and female rats were exposed for 6
hours/day 5 days/week for 13 weeks to MTBE vapors of 0, 800, 4000, and
8000 ppm. During the first 3 weeks of exposure, body-weight gains
decreased. To minimize the influence of the sedation effects associated
with acute exposure, behavioral functioning of the animal was evaluated
after a significant post-exposure period for each test interval; 19
hours for motor activity and 50 hours for FOB. Each time they were
examined (at 4, 8, and 13 weeks), male rats in the group that received
the 4000 ppm MTBE dose showed a decrease in hind-limb grip strength. By
week 8, mean motor activity over 90 minutes was decreased in male rats
in the group that had received the 8000 ppm dose, whereas female rats
that had been given the 4000 ppm MTBE dose exhibited increased activity
at weeks 8 and 13. The transient nature of these effects and the lack
of a dose-response relationship suggest that exposure to MTBE for 13
weeks at concentrations of 8000 ppm or below did not produce
toxicologically significant neurobehavioral changes in the Fischer 344
rat. Significant decreases in absolute mean brain weight were reported
in both male (5%) and female (3%) rats exposed to 8000 ppm MTBE. During
exposure, MTBE produced a significant decrease in body weight gain over
a period of substantial growth. If the brain weight decrease was
evaluated relative to the decrease in body weight gain, the relative
change in brain weight was not different from that of control rats. A
slight decrease in brain length (0.5 mm) was reported to be significant
in male rats exposed to 8000 ppm MTBE. No changes were seen in brain
width and females showed no effect on either brain parameter. Given the
inherent variability associated with excising the brain from the
cranium, and the limit of resolution and accuracy of measurement
techniques, the slight decrease in brain length, while given numerical
significance, is below the level of concern for biological
significance. In the determination of statistical significance for
brain parameters, the lack of adjustment for multiple comparison greatly
influenced the number of values that reached the critical significance
levels were not corrected for the number of comparisons performed thus,
increasing the likelihood of type 1 statistical errors and inflating the
level of significance. Histological examination of various brain
regions and peripheral nerves failed to indicate evidence of
morphological alterations in nervous system tissue after 13 weeks of
exposure (Dodd and Kintigh, 1989).

Robinson and co-workers (1990) examined the oral toxicity of MTBE
and found thatadult rats exhibited profound anesthesia immediately after
oral exposure to MTBE (1428 mg/kg). Within 2 hours, all animals
recovered normal motor and sensory functions. No additional clinical
effects were seen during the 14 days of dosing or during the
90 days when the rats received oral doses of 1200 mg/kg MTBE.

Studies examining the effects of MTBE exposure during gestation
report similar anaesthetic effects of hypoactivity and ataxia in mice
dams at doses of 4000 and 8000 ppm (Bushy Run Research Center, #52-526,
1989). No adverse effects have been reported in rats receiving lower
doses, (API, #32-30235, 1984) or in mice receiving doses up to 2500 ppm
(API, #32-30237, 1984). In rabbits, doses of 4000 and 8000 ppm produced
a decrease in body weight in pregnant dams; however, no other clinical
signs were noted (Bushy Run Research Center #51-628, 1989).

Exposure to MTBE has been reported to transiently affect muscle
creatine-kinase activity with an inhibition early in exposure followed
by an increase in activity by the 15 week of exposure to 300 ppm
(Savolainen et al., 1985). Although levels of MTBE and tertiary-butanol
in the brain were similar in these animals, no effect was seen on brain
succinate dehydrogenase, creatine kinase, or acetylcholinesterase
activities.

3. Chronic Exposure

In a chronic inhalation study (Chun et al., 1992), Fischer 344
rats were exposed toanalytical mean concentrations of MTBE (403, 3023,
or 7977 ppm) 6 hours/day, 5 days/week for 24 months. At high
concentrations, clinical signs of ataxia were evident
starting on the second day. No alterations in brain weight were
observed. Survival time for males was decreased in both the mid- and
high-concentration groups while no decrease was seen in females.
An 18-month inhalation study in CD-1 mice (402, 3014, and 7973 ppm, 6
hours/day, 5 days/week) showed ataxia at the high-exposure level. A 6%
decrease was seen in absolute brain weight in both males and females,
however, this occurred in the presence of a body weight decrease
(Burleigh-Flayer et al., 1992).

The existing data on MTBE neurotoxicity suggests a low potential for
producing neurotoxicity.

4. Teritary-butyl alcohol (TBA)

Although limited research exists on TBA toxicity, reports show similar
effects to those seen with exposure to MTBE. Among humans there are
reports of eye, nose, throat, and skin irritation. Prolonged exposure
to high concentrations of the chemicalcan produce narcotic effects such
as headache, dizziness, drowsiness, and stupor. This
chemical was nominated for testing by the National Toxicology Program
because of itspresence in drinking water. The resulting subchronic
toxicity studies reported clinical
signs in rats of ataxia in both males and females and hypoactivity in
males after the rats drank water containing 4% TBA for 95 days (3588.5
mg/kg/day). In mice, clinical signs of ataxia, abnormal posture, and
hypoactivity were evident (6247 mg/kg/day). These clinical signs are
similar to those for MTBE, could be directly related to chemical
toxicity, and could indicate alcohol inebriation secondary to
dehydration (Lindamood et al. 1992).

5. Summary

All of the animal experimental data suggest that exposure to high
levels of MTBE results in transient sedative effects. The types of
effects reported, ataxia and sedation, support the neuroactive
properties of MTBE. The transient nature of these effects and their
dissipation with a few hours following cessation of exposure
suggest that these neuroactive effects are not associated with
subsequent neurotoxicity. Effects on brain weight varied between
studies with effects seen following 13-18 weeks of exposure in the
presence of decreased body weight growth, while no changes were seen
following acute, 14 day, or 24 month exposure to MTBE. It is therefore
concluded that MTBE has a low potential for producing neurotoxicity.

A. Genetic Toxicity Studies

1. Summary of Genetic Toxicity Studies

a. Oxygenated fuel additives

(1) MTBE

MTBE has been extensively tested for genetic toxicity. In vitro, it
was nottoxic or mutagenic with or without exogenous metabolic activation
in the Salmonella mutation (Ames) test (Cinelli, et al., 1992; Life
Science Research, 1989b; ARCO, 1980; NTP, unpublished), and did not
produce mitotic gene conversion in yeast with or without exogenous
metabolic activation (ARCO, 1980). It was not mutagenic in cultured
Chinese hamster V79 cells (Cinelli, et al., 1992; Life Science Research,
1989a), but was mutagenic in mouse lymphoma L5178Y cells with, but not
without, exogenous metabolic activation (ARCO, 1980). Two different
grades of MTBE, 99% and 99.9% pure, were tested with similar results.
This positive response was explained by the testing laboratory as being
the result of formaldehyde which was produced during the in vitro
metabolism of MTBE and is very active in this test system (ARCO, 1980).

As a result of the positive results with L5178Y cells, MTBE was
tested at similar concentrations for its ability to induce chromosome
aberrations and sister chromatid exchanges (SCE) in Chinese hamster
ovary cells in vitro, with and without exogenous metabolic activation
(ARCO, 1980; Brusick, 1979). Two different grades of MTBE were tested,
99% and 99.9% pure. No induction of chromosome aberrations were seen
with either grade, with or without exogenous metabolic activation (ARCO,
1980; Brusick, 1979). In the SCE test the 99.9% sample showed a
positive response without activation; this result was not repeatable and
the testing laboratory concluded that the overall response was negative
(ARCO, 1980; Brusick, 1979). The 99% sample was negative under all test
conditions (ARCO, 1980; Brusick, 1979).

MTBE did not damage DNA of primary rat hepatocytes in culture as
measured by induction of unscheduled DNA synthesis (Cinelli, et al.,
1992; Life Science Research, 1989c). A structure-activity relationship
analysis resulted in a prediction that MTBE would not be mutagenic in
Salmonella or produce chromosome aberrations or sister chromatid
exchanges in cultured mammalian cells (Rosenkranz and Klopman, 1991).

In vivo, inhalation of MTBE did not produce chromosome aberrations
in bone marrow cells of male or female rats exposed to 0, 800, 4000, and
8000 ppm, 6 hours/day for 5 days (Bushy Run Research Center, 1989), or
micronuclei in the bone marrow cells of male or female mice exposed to
0, 400, 3000, and 8000 ppm, 6 hours/day for 2 days (Bushy Run Research
Center, 1993). Intraperitoneal administration of 0.04, 0.133, or 0.4 ml
MTBE/kg to male rats as single, acute doses, or as 5 consecutive doses,
did not produce an increase in chromosome aberrations in femoral bone
marrow cells (ARCO, 1980). Oral administration of 0, 1, 10, 100, and
1000 mg/kg to male and female CD-1 mice for 3 weeks did not produce
mutations at the hypoxanthine-guanine phosphoribosyl transferase locus
in lymphocytes (Ward et al., 1995). MTBE did not induce sex-linked
recessive lethal mutations in the fruit fly, Drosophila melanogaster,
when administered at 0, 0.03, 0.15, and 0.30% in its food (Hazleton
Laboratories America, Inc., 1989)

(2) ETBE

The only available genetic toxicity data for ETBE is a mutagenicity
test in Salmonella, where no toxicity or mutagenicity was seen (Zeiger
et al., 1992). A structure-activity relationship analysis resulted in a
prediction that ETBE would not be mutagenic in salmonella or produce
chromosome aberrations or sister chromatid exchanges in cultured
mammalian cells (Rosenkranz and Klopman, 1991).

(3) tert-Methyl amyl ether (TAME)

TAME was not mutagenic or toxic in Salmonella when tested with and
without exogenous metabolic activation (Daughtrey and Bird, 1995). It
also did not induce micronuclei in bone marrow cells of male or female
mice injected intraperitoneally with doses of 0, 150, 375, or 750 mg/kg
body weight and sampled at 24, 48, and 72 hours after administration
(Daughtrey and Bird, 1995).

b. Metabolites and photoxidation products

(1) tert-Butyl alcohol

TBA was not toxic or mutagenic in the Salmonella mutagenicity test
(Zeiger et al., 1987), or in the L5178Y mouse lymphoma mutagenicity test
(McGregor et al., 1988). No increases in chromosome aberrations or
sister chromatid exchanges were seen in Chinese hamster ovary cells
treated in culture (National Toxicology Program (NTP), unpublished).

(2) Formaldehyde

Formaldehyde has been extensively tested for genetic
toxicity, and the results have been summarized elsewhere (International
Agency for Research on Cancer [IARC], 1987, 1995). It produced DNA
strand breaks and mutations in bacteria, yeast, and fungi; DNA strand
breaks and cross-links, chromosome aberrations, and sister chromatid
exchanges in cultured human and rodent cells; and sex-linked recessive
lethal mutations and reciprocal translocations in the fruit fly,
Drosophila.

Formaldehyde produced DNA damage in monkey and rat cells treated
in vivo. Mixed results were obtained in other in vivo genetic tests in
rodents and humans (IARC, 1995). Mutations were detected in the p53
gene isolated from nasal tumors of F344 rats following inhalation
exposure to 15 ppm formaldehyde for 2 years (Recio, et al., 1992).
Chromosome aberrations, sister chromatid exchanges, and micronuclei were
induced in rat and mouse cells treated in vivo in some studies but not
in others. Similarly, chromosome aberrations, sister chromatid
exchanges, and micronuclei were found in humans exposed to formaldehyde
in work environments in some studies but not others. Many of these
differences could be the result of different treatment regimens and test
protocols or of differences in the performance of the test
laboratories. Formaldehyde produced sperm-morphology changes in rats
but not in mice or humans (IARC, 1987, 1995).

Most in vivo rodent genetic toxicity studies used only single-dose
treatment. However, a few inhalation studies used multiple dose
treatments. Dallas et al. (1992) exposed male Sprague-Dawley rats to 0,
0.5, 3, or 15 ppm formaldehyde, 6 hours/day, 5 days/week, for 1 or 8
weeks. Dose-related increases in chromosome aberrations were seen in
pulmonary macrophages; the responses at 15 ppm were significantly
increased at both treatment times. There were no chromosome aberrations
produced in bone marrow cells at either treatment time. In a similar
study, Kligerman et al. (1984) treated male and female F344 rats with
formaldehyde by inhalation at 0, 0.5, 6, or 15 ppm for 6 hours/day for 5
days. No increases in sister chromatid exchanges or chromosome
aberrations were seen in peripheral blood lymphocytes.

(3) t-Butyl formate (TBF)

No information was found.

2. Summary of available genetic toxicity data

MTBE has been extensively tested for genetic toxicity with
generally negative results. The only positive result, in an in vitro
test, was attributed to the mutagenicity of the in vitro metabolite,
formaldehyde. Although TAME was tested only in Salmonella and for
micronuclei in mouse bone marrow cells, with negative results, these two
tests are generally considered sufficient to define a nongenotoxic
chemical. There are insufficient data from which to evaluate the
genetic toxicity of ETBE. Similarly, among the metabolites,
formaldehyde has been extensively studied, but there are insufficient
data for TBA. No test data was available for t-butyl formate. Of these
chemicals, the only one with demonstrated, unequivocal mutagenicity is
formaldehyde.

3. Need for genetic-toxicity data

With the exception of MTBE and formaldehyde, and to a more limited
extent with TAME, there are insufficient in vitro or in vivo data
available with which to evaluate the genetic toxicity of these
substances. In vitro and in vivo data are needed for ETBE and TBF, and
in vivo data are needed for TBA.

B. Carcinogenicity Studies

1. MTBE

a. Oral administration in rats

MTBE was tested for carcinogenicity by oral administration in rats at
the Bentivoglio Castle Cancer Research Center in Bologna, Italy.
Exposure to MTBE caused dose-related increases in lymphomas and
leukemias in females and an increase in interstitial cell tumors of the
testes in males (Belpoggi et al., 1995). Individual animal data from
this study are not available.

Groups of 60 male and 60 female Sprague-Dawley rats, 8 weeks of
age, were administered MTBE (> 99 % purity) in olive oil by gavage at
doses of 0 (vehicle control), 250, or 1000 mg/kg body weight, 4 days per
week for 104 weeks. Animals were kept under observation for their
natural lifetime. Survival was higher in the 1000 mg/kg dose group of
males than in control rats and was reduced in both treatment groups of
females. Histopathological evaluations revealed a significant
dose-related increase in the combined incidence of lymphomas and
leukemia in female rats (2/60 in the control rats, 6/60 in the low-dose
group, and 12/60 in the high-dose group), but not in male rats.
Dysplastic proliferation of lymphoreticular tissue was also increased in
treated female rats (1 control rat, 15 in the low-dose group, and 9 in
the high-dose group). It should be noted that McConnell et al. (1986)
recommended against combining mononuclear cell leukemias (MNCL) with
lymphomas in F344 rats. In that strain of rat, MNCL occurs at a high
spontaneous rate. For Sprague-Dawley rats, however, it is probably
reasonable to combine leukemia and lymphoma, especially when the
leukemias are of lymphocytic origin. Interstitial cell (Leydig cell)
tumors of the testes were significantly increased in the high dose group
of male rats (2/60 control rats, 2/60 in the low-dose group, and 11/60
in the high dose group). An increase in uterine sarcomas was observed
in the low dose group of female rats but not in the high dose group (1
control rat, 5 in the low-dose group, 0 in the high-dose group).
Decreases in the incidence of fibromas and fibroadenomas of the mammary
gland (40 in the control group, 27 in the low-dose group, 16 in the
high-dose group), pituitary gland adenomas (22 in the control group, 16
in the low-dose group, and 13 in the high-dose group), and adrenal gland
pheochromocytomas (18 in the control group, 11 in the low-dose group,
and 10 in the high-dose group) in treated female rats may have been due
to increased early mortality in the dosed groups.

b. Inhalation exposure in mice

MTBE was tested for carcinogenicity in mice by inhalation exposure
at the Bushy Run Research Center, Export, Pennsylvania. Exposure to 8000
ppm MTBE produced increases in hepatocellular carcinomas in male mice
and in hepatocellular adenomas in female mice (Burleigh-Flayer et al.,
1992).

Groups of 50 male and 50 female CD-1 mice, 6 to 7 weeks of age, were
exposed to filtered air containing 0 (control), 400 ppm, 3000 ppm, or
8000 ppm MTBE (99% purity) [0, 1,440, 10,800, or 28,800 mg/m3],
6 hours/day, 5 days/week, for 18 months. Mortality was increased, and
mean survival time was decreased only in the high-exposure group of male
mice. Mean survival time was 510 days for control male mice and 438
days for high-exposure male mice. An increased incidence of
hepatocellular carcinomas was observed in the
8000 ppm exposure group of males (2/49 control mice, 4/50 in the low-
exposure group, 3/50 in the mid-exposure group, and 8/49 in the
high-exposure group). An increased incidence of hepatocellular adenomas
and carcinomas in the high-exposure group of males compared with the
incidence in the control group was not significant by the Fisher's Exact
Test (12 control mice, 12 mice in the low-exposure group, 12 in the
mid-exposure group, and 16 in the high-exposure group); however, this
method of analysis does not adjust for differences in survival between
the control and exposure groups. In female mice, exposure to 8000 ppm
MTBE produced a significant increase in the incidence of hepatocellular
adenomas and carcinomas (2/50 in the control group, 2/50 in the
low-exposure group, 2/50 in the mid-exposure group, and 11/50 in the
high-exposure group). Liver tumors were predominantly adenomas but did
include one hepatocellular carcinoma in the 400 ppm exposure group and
one in the 8000 ppm exposure group. A significant decrease in the
incidence of hepatocellular degeneration was observed in the high
exposure group of female mice (34 in the control group, 24 in the
low-exposure group, 25 in the mid-exposure group, and 17 in the
high-exposure group). Because the duration of this study was only 18
months, it is not possible to determine whether MTBE induces
late-developing tumors in the low- or mid-exposure groups or at other
sites in mice, or whether more hepatocellular adenomas would have
progressed to carcinomas. The National Toxicology Program (NTP) uses
2-year treatment periods to evaluate the chronic toxicity and
carcinogenic potential of chemicals in mice and rats.

c. Inhalation exposure in rats

MTBE was tested for carcinogenicity in rats by inhalation exposure
at the Bushy Run Research Center. Exposure to MTBE produced increases
in uncommon kidney tumors and commonly occurring interstitial cell
tumors of the testes in male rats (Chun et al., 1992).

Groups of 50 male and 50 female F344 rats, 6 to 7 weeks of age,
were exposed tofiltered air containing 0 (control), 400 ppm, 3000 ppm,
or 8000 ppm MTBE (99% purity),
6 hours/day, 5 days/week, for 2 years. Because of the increased
mortality of males exposed to 3000 and 8000 ppm, surviving rats in these
exposure groups were killed after 97 weeks and 82 weeks of exposure,
respectively. Mean survival time of male rats was significantly reduced
in all of the exposure groups compared with that of rats in the control
group (632 days for control rats, 617 days for the low-exposure group,
587 days for the mid-exposure group, and 516 days for the high-exposure
group). Body weights of males and females in the 8000 ppm exposure
group were less than those of control rats. Exposure-related increases
in the severity of chronic nephropathy were observed in all exposure
groups of males and in the 3000 and 8000 ppm exposure groups
of females. Nephropathy is an age-related disease characterized by
degeneration and atrophy of the tubular epithelium, dilation of tubules
with the formation of hyalin and granular casts, regeneration of tubular
epithelium, glomerulosclerosis, and interstitial inflammation and
fibrosis. In a previous 13-week inhalation study of MTBE in F344 rats
at exposures of 0, 800, 4000, or 8000 ppm (Dodd and Kintigh, 1989), an
increase in hyaline droplet formation was observed in the renal proximal
tubules of male rats in the high-exposure group; however, this change
was not accompanied by other lesions characteristic of alpha-2u-globulin
nephropathy, including degeneration and regeneration of the renal
tubular epithelium and accumulation of granular casts within tubular lumens.

Chronic progressive nephropathy was considered to be a main cause of
the early deaths of males in the 3000 and 8000 ppm exposure groups.
Immunohistochemical staining of kidney sections for alpha-2u-globulin in
F344 rats exposed to 0 (control), 800, 4000, or
8000 ppm MTBE for 13 weeks did not provide evidence of an
exposure-related increase in the intensity or staining area of this
protein (percentage of cortex staining for alpha-2u-globulin: 24%
control, 48% low-exposure, 36% mid-exposure, and 44% high exposure)
(Swenberg and Dietrich, 1991). Furthermore, proteinaceous casts
localized at the junction of the proximal tubules and the thin loop of
Henle did not stain positively for alpha-2u-globulin. These results
indicate that accumulation of alpha-2u-globulin was not causally related
to the elevated mortality in male rats resulting from increased severity
of chronic nephropathy in the MTBE exposure groups. Alpha-2u-globulin
is a low molecular weight protein synthesized in the liver of male rats
but not by hepatocytes of female rats or by mice of either sex. The
finding that the severity of chronic nephropathy was also increased in
female rats exposed to MTBE does not support the hypothesis that
alpha-2u-globulin alone accounted for the kidney lesions observed in
these studies.

Incidences of renal tubule adenomas and carcinomas were increased
in exposed males (one control rat, none in the low-exposure group, eight
in the mid-exposure group, and three in the high-exposure group). Three
of the eight kidney tumors in the 3000 ppm exposure group were renal
tubule carcinomas. No carcinomas were observed in the other treatment
groups or in the control group. In addition, preneoplastic adenomatous
tubular cell hyperplasia was diagnosed in the kidney of two male rats in
the low-exposure group. The latter change represents an early stage in
the continuum of renal lesions that can progress from atypical
hyperplasia to adenoma to carcinoma (Hard, 1986). One renal tubule
adenoma was detected in the mid-exposure group of females. An
exposure-related increase in incidence of interstitial cell adenomas of
the testes was also observed (32/50 in the control group, 35/50 in the
low-exposure group, 41/50 in the mid-exposure group, and 47/50 in the
high-exposure group). These tumors occur at a high spontaneous rate in
this strain of rat. The level of significance of testicular tumors
should be reanalyzed using statistical methods that adjust for
differences in survival between the control and exposure groups.

2. Carcinogenicity of MTBE Metabolites

a. t-Butyl Alcohol (TBA)

TBA, a metabolite of MTBE, was tested for carcinogenicity (NTP,
1995) in rats and in mice by administration in drinking water. Exposure
to TBA produced increases in uncommonly occurring kidney tumors in male
rats and follicular cell adenomas of the thyroid gland in female mice
(Cirvello et al., 1995; NTP, 1995).

(1) Rats

99% purity) in deionized drinking water at concentrations of 0, 1.25,
2.5, or 5 mg/ml for up to 103 weeks while groups of 60 female F344 rats
were given TBA in their drinking water at concentrations of 0, 2.5, 5,
or 10 mg/ml. Ten male and female rats from each group were killed after
15 months of the study. The average delivered daily doses of TBA were
approximately 85, 195, and 420 mg/kg body weight for males and 175, 330,
and 650 mg/kg for females.

At the 15-month evaluation, dose-related increases in kidney weights
were observedin males and females. Nephropathy was present in all
control and exposed rats but was
generally more severe in the treatment groups. A renal tubule adenoma
was observed in one male rat from the high-dose group. At the end of
the 2-year study, survival was significantly lower for the males in
high-dose group (1/50) and females in the high-dose group (12/50) than
for rats in the control groups (10/50 for males and 28/50 for
females). Incidences of focal renal tubule hyperplasia and renal
tubule adenoma were increased in exposed males. Hyperplasia was found
in 3 of 50 controls rats, 7 of 50 in the low-dose group, 6 of 50 in the
mid-dose group, and 6 of 50 in the high-dose group. Renal tubule
adenomas were found in one control male rat, in three in the low-dose
group, in four in the mid-dose group, and in three in the high-dose group.

In the high-dose group, a renal tubule carcinoma was observed in a male
rat that also had a renal tubule adenoma. The severity of nephropathy
was increased significantly in males in the high-dose group and in all
of the exposure groups of females. Renal tubule hyperplasia was seen in
one female rat in the high dose group. Because of the observed increase
in uncommon proliferative lesions of the renal tubules in exposed males,
six to eight additional sections per kidney from all control and treated
males were prepared and examined microscopically. The extended
evaluation revealed significant increases in the incidence of renal
tubule hyperplasia in the high-dose group and of renal tubule adenoma in
the mid-dose group. Renal tubule hyperplasia for the standard and
extended evaluations combined was found in 14 control rats, 20 rats in
the low-dose group, 17 in the mid-dose group, and 25 in the high-dose
group; renal tubule adenomas or carcinomas were found in 8 control rats,
13 rats in the low-dose group, 19 rats in the mid-dose group, and 13
rats in the high-dose group. Four renal tubule carcinomas were seen in
this study; two were found in the low-dose group, one in the mid-dose
group, and one in the high-dose group.

In the preliminary 13-week drinking water study of TBA in rats (at
concentrations up to 40 mg/ml), a treatment-related increase in hyaline
droplet accumulation was observed in the renal proximal tubules of male
rats but not in female rats. In addition, an increase in renal tubular
cell replication was observed in male rats, but only at exposures that
exceeded the carcinogenic doses used in the 2-year study of TBA
(Takahashi et al., 1993).

(2) Mice

Groups of 60 male and 60 female B6C3F1 mice, 7 weeks of age were
given TBA (> 99% purity) in deionized drinking water at concentrations
of 0, 5, 10, or 20 mg/ml for 103 weeks. The average delivered daily
doses of TBA were approximately 535, 1035, and 2065 mg/kg body weight
for males and 510, 1015, and 2105 mg/kg for females. Survival of males
in the high-dose group was significantly lower (17/60) than that of the
control group (27/60).

The incidences of follicular cell hyperplasia and adenoma of the
thyroid gland were increased in treated males. Hyperplasia was found in
5 of 60 control male mice, in 18 of 59 the low-dose group, in 15 of 59
in the mid-dose group, and in 18 of 57 in the high-dose group. Adenomas
were found in 1 control male mouse, in no mice in the low dose group, in
4 of 59 in the mid-dose group, and in 1 in the high-dose group. In
treated females, hyperplasia was seen in 19 of 58 control mice, 28 of 60
in the low dose-group, 33 of 59 in the mid-dose group, and 47 of 59 in
the high-dose group. Adenomas were found in two control female mice,
three mice in the low-dose group, two in the mid-dose group, and nine in
the high-dose group. In addition, a follicular cell carcinoma was
present in one male mouse in the high-dose group. The tumor response in
female mice in the high-dose group was statistically significant.

Incidences of chronic inflammation and transitional
epithelial hyperplasia of theurinary bladder were increased in males and
females in the high-dose groups compared with the control groups. There
was no evidence of progression of these lesions to urinary bladder
neoplasia.

b. Formaldehyde

On the basis of animal and human data, an expert panel of the IARC
concluded that formaldehyde is probably carcinogenic to humans (IARC,
1995).

(1) Inhalation

Inhalation exposure of male and female B6C3F1 mice to 0, 2.0, 5.6,
or 14.3 ppm (0, 2.5, 6.9, 17.6 mg/m3) formaldehyde for 6
hours/day, 5 days/week, for 24 months, followed by a 6-month observation
period with no further exposure, did not produce a carcinogenic response
(Kerns et al., 1983).

Squamous cell carcinomas of the nasal cavity were induced in male
and female F344 rats exposed to 14.3 ppm formaldehyde for 24 months (in
males: 0/118 control rats, 0/118 at 2.0 ppm, 1/119 at 5.6 ppm, and
51/117 at 14.3 ppm; in females: 0/114 control rats, 0/118 at 2.0 ppm,
0/116 at 5.6 ppm, and 52/115 at 14.3 ppm) (Kerns et al., 1983). Five
additional nasal cavity tumors were observed in rats exposed to 14.3
ppm. In a lifetime inhalation study, exposure of 100 male
Sprague-Dawley rats to 14.3 ppm formaldehyde (6 hours/day, 5 days/week)
produced 38 squamous cell carcinomas of the nasal cavity and 10 polyps
or papillomas; neither of these lesions was observed in control rats
(Sellakumar et al., 1985). There were no increases in tumor incidence
beyond the respiratory tract. Exposure of Wistar rats to 10 ppm
formaldehyde for 28 months (6 hours/day, 5 days/week) produced an
increased incidence of nasal tumors in rats in which the nasal mucosa
had been severely injured by bilateral intranasal electrocoagulation
(17/58 versus 1/54 in control rats with damaged noses); in rats with
undamaged noses, the incidence of nasal tumors was 1/26 in rats exposed
to 10 ppm formaldehyde and 0/26 in control rats (Woutersen et al., 1989).

Exposure of male Syrian golden hamsters to 10 ppm formaldehyde (5
hours/day, 5 days/week for life) or to 30 ppm formaldehyde (for 5 hours
once per week for life) did not induce tumors of the nasal cavity or
respiratory tract (Dalbey, 1982).

(2) Oral administration

In a lifetime study in male and female Sprague-Dawley rats,
formaldehyde was administered in drinking water at concentrations of 0,
10, 50, 100, 500, 1000, or 1500 ppm (Soffritti et al., 1989). A
dose-related increased incidence of leukemia was observed (7/200 in
control rats and 3/100, 9/100, 9/100 12/100, 13/100, and 18/100 in rats
in the respective dose groups). Intestinal tumors were observed in
6/100 rats in the high-dose group but not in any rats in the control
group. In a 24-month drinking water study in Wistar rats, doses of
formaldehyde up to 82 mg/kg body weight for males and 109 mg/kg for
females did not produce any treatment-related carcinogenic response (Til
et al., 1989).

3. Cancer Potency (see Table 5)

A full description of the cancer risk derivations for MTBE is
presented in Appendix A. Based on the liver tumor data from the 18
month inhalation study of MTBE in male and female CD-1 mice
(Burleigh-Flayer et al., 1992), an estimate of the upper-bound unit
cancer risk for lifetime human inhalation exposure to MTBE using the
linearized multistage model is 6 x 10-4 per ppm (2 x
10-7 per ug/m3).
The human cancer ED10 benchmark dose (i.e., the estimated effective dose
associated with an increased cancer risk of 10%) was estimated to be 460
ppm based on the mouse liver tumor data.

On the basis of the kidney tumor data from the 2-year inhalation
study of MTBE in male F344 rats (Chun et al., 1992), the estimated
upper-bound unit cancer risk for lifetime inhalation exposure to MTBE
using the linearized multistage model is 6 x 10-4 per ppm (2 x
10-7 per ug/m3) and the human cancer ED10
benchmark dose is 330 ppm. Because of
reduced survival in the male rat exposure groups, upper bound unit
cancer risk estimates based on the kidney tumor data were calculated
using survival-adjusted tumor rates. Even with this adjustment, the
highest exposure group (8000 ppm) had to be excluded from the analysis
because it failed a goodness-of-fit test. Eliminating this group from
the analysis is acceptable for low dose risk estimations because the
mortality was greatest in this exposure group, the tumor response in
this group was not significant (probably because of early mortality),
and the mid exposure group (3000 ppm) had a significant tumor
response. Preliminary estimates of cancer potency had been made using
a Weibull time-to-tumor model to account for differences in survival.
However, because predictions of tumor incidence from this model were
inconsistent with the experimental data from the lowest exposure group,
use of this model was discontinued in favor of the linearized multistage
model.

The inhalation upper bound cancer unit risks of MTBE are
slightly lower than those of fully vaporized conventional gasoline which
has been listed by EPA as a probable human carcinogen based on animal
carcinogenicity data. In both cases (MTBE and fully vaporized
conventional gasoline), estimates of unit cancer risk were made using
the linearized multistage model and based on the induction of liver
tumors in mice and kidney tumors in rats from inhalation exposures. If
the lymphoma/leukemia data from the gavage study are used to estimate
unit cancer risk for inhalation exposure (i.e., assuming equivalent
total exposures by inhalation and gavage result in similar internal
doses of MTBE and that differences in dose rate do not affect the tumor
response), then the estimated upper bound unit cancer risk (4 x
10-3 per ppm) is similar to that of fully vaporized
conventional gasoline. It should be noted, however, that human
exposures to nonoxygenated
gasoline vapors may be very different than exposures to fully vaporized
conventional gasoline. The estimated upper bound unit cancer risk for
MTBE using the lymphoma/leukemia data is approximately an order of
magnitude lower than that of benzene, a constituent of gasoline that is
classified as a known human carcinogen. The unit cancer risk of MTBE was
more than 100 times less than that of 1,3-butadiene, a carcinogenic
emission product of incomplete fuel combustion.

4. Summary of Animal Carcinogenicity Data and Overall
Evaluation

Experimental studies demonstrate that MTBE is carcinogenic in animals
by either the oral or inhalation routes of exposure, with tumor
responses seen at multiple organ sites. TBA and formaldehyde, the
primary metabolites of MTBE, are also carcinogenic in animals exposed
directly to these substances. Inhalation exposure to MTBE produced
increased incidences of tumors of the kidney and testes in male rats and
of liver tumors in mice. Oral administration of MTBE produced
lymphomas and leukemia in female rats and tumors of the testes in male
rats. Oral administration of TBA produced tumors of the kidney in male
rats and tumors of the thyroid gland in female mice. Formaldehyde
produced tumors of the nasal cavity after inhalation exposure and leukemia
after oral administration in rats. There is no justification to exclude
any of these responses for hazard identification nor is there sufficient
evidence to support deviating from default linear models for estimations
of human risk.

The experimental studies provide sufficient evidence for the
carcinogenicity of MTBE in animals. There are no studies on the
carcinogenicity of MTBE in humans. Using IARC criteria for evaluating
evidence of carcinogenicity, MTBE could be classified as possibly
carcinogenic to humans (Group 2B); using EPA criteria of 1986, MTBE
could be classified as a probable human carcinogen (Group B2). Other
data (e.g., tissue dosimetry of formaldehyde resulting from exposure to
MTBE) could influence the overall evaluation. Epidemiologic studies
suggest a causal relationship between exposure to formaldehyde and
nasopharyngeal cancer (IARC, 1995). Because all known human carcinogens
that have been adequately tested produced positive carcinogenic results
in experimental animals, it is plausible and prudent to regard agents
that have sufficent evidence of carcinogenicity in animals as presenting
a carcinogenic risk to humans (IARC, 1995). The existing data on MTBE
indicate that this chemical should be regarded as posing a potential
carcinogenic risk to humans, while recognizing that the estimated upper
bound cancer unit risks of MTBE are similar to or slightly less than
those of fully vaporized conventional gasoline.

No studies have been reported on the carcinogenicity of ETBE,
TAME, or TBF.

Table 5. Cancer Potency Estimates for MTBE Based on
Tumor Data from Studies in Rats and Micea

b For comparison, estimated upper bound unit cancer risks for
fully
vaporized conventional gasoline are noted: 2 x 10-3 per ppm
based on induction of liver tumors in mice and 4 x 10-3 per ppm
based on induction of kidney tumors in rats (EPA, 1987).

C. Reproductive and Developmental Effects

Published and unpublished reports on the effects of MTBE on mammalian
reproduction and development were reviewed. Those studies are
summarized below. No such studies of ETBE or TAME were located.

1. Developmental Toxicology Studies of MTBE

Tests for adverse developmental effects of MTBE have been
conducted in mice, rats, and rabbits. From studies involving 1000, 4000,
and 8000 ppm exposures of mice and rabbits, no-observed-adverse-effect
levels (NOAELs) were reported as 1000 ppm for maternal and developmental
toxicity in mice and 8000 ppm for teratogenicity and 1000 ppm for
maternal toxicity in rabbits and 8000 ppm for developmental toxicity.
No adverse maternal or fetal effects were observed in a rat study
involving doses of 250, 1000, and 2500 ppm.

a. Mice

Conaway et al. (1985) exposed pregnant CD1 female mice to MTBE via
inhalation at target levels of 0, 250, 1000, or 2500 ppm on gestation
days (Gds) 6 through 15. Food and water consumption and maternal weights
were recorded throughout dosing. Females
were killed on Gd 18, and their uterine contents evaluated for numbers
of implantations, number and weight of live and malformed fetuses, and
number of late and early resorptions. Each fetus was weighed, its sex
determined, and its length measured. One-third of the fetuses were
examined for soft tissue malformations via Wilson's free-hand sectioning
technique, and the remaining two-thirds were evaluated for skeletal
malformations after Alizarin staining. No maternal toxicity was
observed in this study, nor were there any significant changes in fetal
weights. A small but significant decrease in the number of viable
implants was observed at 250 ppm and 2500 ppm but not at 1000 ppm. This
decrease in viable implants in the low-and high-dose groups was
attributed to two females in each group that had an unusually high
number of resorptions. No significant increase in fetal malformations
was reported.

A mouse study conducted at the Bushy Run Research Center (Bushy
Run Research Center, Project Report No. 52-526, 1989; Neeper-Bradley,
1990) involved higher levels of MTBE than those in the Conaway et al.
study. Pregnant CD1 females were exposed via inhalation to target-dose
levels of 0, 1000, 4000, and 8000 ppm on Gds 6-15. Food and water
consumption and maternal weights were recorded throughout dosing.
Maternal toxicity was observed at 4000 and 8000 ppm with decreased food
consumption, reductions in body weight and weight gain, labored
respiration, and lacrimation with other signs of toxicity observed at
8000 ppm and hypoactivity and ataxia observed at 4000 and 8000 ppm. The
uterine contents of females killed on Gd 18 were evaluated for number
of implantation sites, early and late resorptions, and dead and live
fetuses. Ovarian corpus luteum counts were also made. All fetuses were
weighed and examined for external malformations. Approximately one-half
of the live fetuses were examined for thoracic-abdominal and cranial
malformations using modified methods of Staples or Wilson,
respectively. All fetuses were processed for skeletal staining with
Alizarin Red. Fetal weight changes and skeletal variations were
observed at 4000 and 8000 ppm, but they were attributed to maternal
toxicities. An increased incidence of cleft palate was observed at 8000
ppm, and the possibility that this condition might also be related to
maternal toxicity was discussed. The number of viable implants was signific
antly decreased only at 8000 ppm. This decrease was due to an increase
in the number of late resorptions and dead fetuses rather than to early
resorptions.

b. Rats

There was one MTBE developmental toxicity study with rats. Conaway
et al. (1985) exposed pregnant Sprague-Dawley females to MTBE via
inhalation at target levels of 0,
250, 1000 or 2500 ppm on Gd 6 through Gd 15. Food and water consumption
and maternal weights were recorded throughout dosing. Females were
killed on Gd 20 and their uterine contents evaluated for numbers of
implantations, number and weight of live or malformed fetuses, and
number of late and early resorptions. Each fetus was weighed, its sex
determined, and its length measured. One-third of the fetuses were
examined for soft tissue malformations via Wilson's free-hand sectioning
technique, and the rem aining two-thirds were evaluated for skeletal
malformations after Alizarin staining preparations. No maternal
toxicity was observed in this study, and there were no significant
changes in fetal weights. There were no significant effects of MTBE on
any of the measured parameters except for a reduction in maternal food
consumption during the Gd 9-12 interval.

c. Rabbits

One MTBE developmental toxicology study with rabbits conducted at the
Bushy Run Research Center (Bushy Run Research Center, Project Report No.
51-628, 1989; Neeper-Bradley, 1990) involved inhalation exposure of
pregnant New Zealand white females to target-dose levels of 0, 1000,
4000, and 8000 ppm on Gds 6-18. Food and water consumption and maternal
weights were recorded throughout dosing. Maternal toxicity was observed
at 4000 and 8000 ppm as significant reductions in weight gain and food
consumption and increased liver weight at 8000 ppm. The uterine
contents of females killed on Gd 29 were evaluated for the number the of
implantation sites, early and late resorptions, and dead and live
fetuses. Ovarian corpus luteum counts were also made. All fetuses were
weighed and examined for external malformations.

Approximately one-half of the live fetuses were examined for
thoracic-abdominal and cranial malformations using modified methods of
Staples or Wilson, respectively. All fetuses were processed for
skeletal staining with Alizarin Red. There were no significant changes
in the number of corpus luteums, implants, resorptions, and dead or live
fetuses, or in the incidence of malformations.

2. Reproductive Toxicology Studies of MTBE

Neeper-Bradley et al. (1991) conducted a two-generation rat study by
inhalation, using vapor-target concentrations of 400, 3000, and 8000
ppm, 6 hours/day, 5 days/week. The 400 ppm dose was a
no-detectable-adverse-effect level. At 3000 ppm, there was a slight
increase in relative liver weights in adult rats at necropsy and a
slight reduction in adult weight gain, as well as slight reductions in
pup weight gain postnatally for both generations. At 8000 ppm, there
were more severe reductions in pup weight gain postnatally (by 10%) in
both generations, and a four-fold increase in pup deaths between birth
and litter standardization at post natal day four. Sires and dams in F0
and F1 generations showed reduced weight gain during exposure, transient
reductions in food intake and increases in adverse clinical signs, and
increased liver weights at F1 adult necropsy. There were no significant
histologic changes noted in any generation. Thus, MTBE was a
reproductive toxicant only at concentrations that also reduced adult
weight gain and increased relative liver weight.

Biles et al. (1987) report a single-generation inhalation study
of MTBE in rats. Exposures were at 0, 300, 1300, and 3400 ppm, 6
hours/day, 5 days/week. Males were exposed for 12 weeks, females for 3
weeks, and then mated while being exposed. These FO rates generated two
litters: both were nursed until 21 days after birth, then killed, and
the pups examined. MTBE exposure did not change any endpoint related to
fertility or mating in the FO rats. There was a 4% reduction in pup
viability in the second litter in the top two dose groups (reduced from
a control value of 99% to 95% for both groups); this reduction is not
considered to be of biological significance. Pup post-mortem
evaluations found no differences between groups. There were no body
weight differences between the groups of FO rats, a finding which
suggesting that a maximum tolerated dose was not reached. No
differences between the groups of FO rats were noticed at necropsy or
microscopically or at necropsy. No sperm measures were taken, nor was
vaginal cyclicity evaluated. In sum, this study found no change in body
weight, and no effects in any reproductive or fertility endpoint were
measured.

3. Reproductive Toxicology Studies of Inhaled Ethanol

Nelson et al. (1985) conducted a rat teratology study in which
pregnant dams were exposed by inhalation to target concentrations of
ethanol vapors of 10,000, 16,000, and 20,000 ppm on Gds 1-19. Dams were
killed on Gd 20. Standard soft tissue and skeletal assessments were
conducted on the fetuses. Researchers reported that at the highest
exposure, dams were severely narcotized and that their food consumption
was reduced by approximately 30% but that maternal body weights were not
affected. Maternal body weights and clinical signs were reported to be
unchanged at the lower exposures. In the high-exposure group, male
fetal weights were reduced by about 8%. No increase in skeletal or soft
tissue malformations were found at any exposure level. There were no
significant exposure-related changes in any malformation or variation.
Thus, at three exposure levels, the highest of which narcotized the
dams, inhaled ethanol caused no adverse affects on the fetuses. However,
it should be noted that the rat has been considered a poor animal model
for ethanol-induced developmental abnormalities.

4. Reproductive Toxicology of Inhaled TBA

Tertiary butyl alcohol (TBA) was assessed in the Sperm Morphology and
Vaginal Cytology Evaluation by the NTP (unpublished, see NTP Technical
Report number 436). F344 male and female rats and B6C3F1 male and
female mice were exposed to TBA by inhalation for 90 days at target
concentrations of 450, 900, and 1750 ppm. In male rats and mice,
exposure led to no changes in caudal, epididymal or testis weights,
sperm motility, sperm density, or percent-abnormal sperm. In female
rats and mice, no changes in estrual cyclicity or average estrous
lengths were observed after TBA exposure. These results suggest that
TBA, under the conditions of this study, did not induce structural
changes in the males or hormonal changes in the females. Determinations
of reproductive function were not made.

5. Summary

Some reproductive and developmental effects have been observed
in rodent models used to assess the health effects of exposure to MTBE.
These reported effects have led EPA to consider MTBE a developmental
toxicant. These limited effects were observed only at exposure levels
that resulted in toxicity to the parental animals. In a preliminary
evaluation, EPA used these results to derive a developmental-toxicity
reference concentration for MTBE of 13 ppm. The
duration of exposure necessary to induce developmental effects is not
known, but negative results have been reported in studies involving
multi-day exposures to high concentrations of MTBE (e.g., 2500 ppm).
Considering the magnitude and duration of exposures in the animal
studies and the association of developmental effects with maternal
toxicity, it is concluded that MTBE is not expected to pose a
reproductive or developmental health hazard under the intermittent,
low-level exposures experienced by humans. No published reports were
located on the reproductive or developmental toxicity of TBA. Long-term
(90-day) inhalation exposure of rats and mice to TBA gave no evidence of
testicular or sperm effects in males or of changes in the estrous cycles
of females. No studies of ETBE or TAME on mammalian reproduction or
development were located.

D. Other Toxic Effects of Long-term Exposure to MTBE

The U.S. EPA has established an inhalation reference
concentration (RfC) of 3 mg/m3 (0.83 ppm) for MTBE (IRIS, 1993).
The
RFC is defined as an estimate (with uncertainty spanning about an order
of magnitude) of a continuous inhalation exposure level for the human
population (including sensitive subpopulations) that is likely to be
without appreciable risk for deleterious noncancer effects during a
lifetime. The MTBE RfC was based primarily on findings of increased
absolute and relative liver and kidney weights and increased severity of
spontaneous renal lesions in female rats, increased prostration in
female rats, and swollen periocular tissue in both male and female rats
as reported in a chronic inhalation study by Chun et al. (1992). These
effects were seen after exposure to 3023 ppm MTBE vapor for 24 months.
Thus, 3023 ppm was a lowest-observed-adverse-effect level (LOAEL), and
403 ppm was a NOAEL for chronic MTBE exposure. Adjusting the NOAEL for
the equivalent human exposure concentration and dividing by a total
uncertainty factor of 100 (for extrapolation from animals to humans, for
protection of sensitive subpopulations, and for data base deficiencies)
yielded an RfC value of 3 mg/m3 (0.83 ppm) MTBE.

Gasoline as an automotive fuel presents both acute and chronic health
hazards. Any consideration of the additive risks of MTBE must be
compared with the risks associated with the gasoline mixtures it
replaces. It is expected that such risk comparisons, although beyond
the scope of this assessment, would be evaluated in the second phase of
the larger interagency assessment of oxygenated fuels that is now under way.

A. Acute Health Effects

Reports of health complaints of various nonspecific symptoms, such
as headache, nausea, and cough, have been reported in various areas of
the country since the introduction of the oxygenated gasoline program
(primarily oxygenated gasoline containing MTBE). The scientific
explanation for these reports of symptoms remains to be determined. In
animal studies, MTBE does not appear to be neurotoxic, and controlled
clinical studies of exposure to pure MTBE among healthy human volunteers
did not measure increased symptoms or measurable indications of
irritation. However, the health effects of MTBE in gasoline may be
different from those of pure MTBE. In addition, the responses of people
who may be more sensitive to MTBE in gasoline might not be adequately
reflected by experimental animals or healthy volunteers. Epidemiologic
studies have been limited, and have neither demonstrated nor ruled out a
causal association between acute health effects and exposure to MTBE or
other oxygenates in gasoline in a relatively small proportion of
persons. Thus, the available scientific evidence regarding exposure to
oxygenated gasoline and acute health symptoms was considered
insufficient to develop estimates of effect at different exposure
levels. Additional controlled exposure studies of people with
self-described sensitivity to oxyfuels, as well as epidemiologic studies
of populations exposed to oxygenated gasoline, are needed to better
understand the risk factors for the reported increase in acute health
symptoms after exposure to oxygenated gasoline.

B. Cancer Risk Characterization

The potential cancer risk of MTBE for humans is best addressed
through an examination of long term epidemiologic studies or long- or
short-term experimental animal studies. Numerous schemes for weighing
evidence are in use today; IARC, EPA, and NTP have established similar
guidelines for considering evidence.

For MTBE, there are no studies of long-term human exposure, a
fact which is not surprising since the use of MTBE is fairly recent.
There are experimental studies in which animals were exposed to high
MTBE doses in order to optimize the ability to detect a carcinogenic
response. For inhalation exposure, both sexes of rats and mice were
tested, and kidney tumors and testicular tumors were seen in male rats
and liver tumors in male and female mice. For gavage exposure, MTBE
caused leukemia or lymphoma in female rats and testicular tumors in male
rats. Two metabolites of MTBE, TBA and formaldehyde, have also been
tested in animals and show carcinogenic activity, with some responses
paralleling those seen with MTBE and some responses being at different
sites. The manner in which MTBE causes tumors to arise is not known.
Increased tumor incidences were not observed at low doses, possibly
because of the small group sizes used in the experiments. Such results
are not sufficient to rule out a possible low dose hazard, given our
lack of understanding about how MTBE causes tumors in animals.

The positive animal studies raise the concern that MTBE may pose
a human health hazard. In the absence of good epidemiologic evidence or
knowledge about how the agent likely reacts in humans, the positive
animal evidence (in two species by two routes of exposure) and other
supporting factors provide evidence that MTBE could pose a human health
hazard under some conditions of exposure. Currently, a low-dose hazard
can not be ruled out. Based on animal data, MTBE is considered to be
either a possible or probable human carcinogen.

It may be useful to ask the "what if" question, i.e., if MTBE is a human
carcinogen, what might be the impact on an exposed human population?
While we have only established that MTBE has potential to be a human
hazard, quantitative human risk estimates can be developed if we make
the assumption that it is a human carcinogen. Public health
conservative assumptions are used to formulate these risk estimates,
which are by design almost always high enough so as to not likely
underestimate the true risk. In fact, the true risks, which we are
rarely able to identify, may be lower than the risk estimates and could
even be negligible.

Estimations of cancer risk for populations exposed to MTBE are based
on estimates of lifetime human exposure, on extrapolation models, and on
several estimates of cancer potency for this chemical. Human exposures
are estimated from measurements of MTBE in ambient air, from
measurements of MTBE concentrations in specific microenvironments (e.g.,
inside automobiles), and from estimates of the distribution of time
spent in each environment. A reasonable worst case exposure scenario
for a typical motorist in the general population that uses oxygenated
fuels containing MTBE might result in annual time-weighted average
exposures of 0.014 ppm MTBE, based on a 4-month oxygenated fuel season,
0.019 ppm based on a 6-month oxygenated fuel season, or 0.029 ppm based
on 6 months of use of oxygenated gasoline and 6 months of use of
reformulated gasoline. These exposure scenarios do not necessarily
represent the worst possible case, but are expected to be well above the
average exposure for the entire population of MTBE-exposed motorists.
The largest occupational group with significant potential exposure to
MTBE is service station attendants; their estimated time-weighted
average exposure to MTBE, including an 8-hour time-weighted average
occupational exposure (5 days per week for 40 years working lifetime) of
0.6 ppm during the 6-month oxygenated fuel season and 0.44 ppm during
the 6-month reformulated gasoline season, is 0.10 ppm.

Cancer potency estimates for MTBE were made based on mouse liver
tumor data from inhalation exposure, rat kidney tumor data from
inhalation exposure, and lymphoma/leukemia data from oral exposure
(Table 5). Alpha-2u-globulin nephropathy was not considered to be
applicable to the kidney tumor response in male rats because:

(i) Other than hyaline droplet formation, renal lesions
characteristic of alpha-2u-globulin nephropathy were not detected in
male rats exposed to MTBE for 13 weeks at concentrations up to 8000 ppm.

(ii) Immunohistochemical staining of kidney sections for
alpha-2u-globulin in male rats exposed to MTBE for 13 weeks at
concentrations up to 8000 ppm did not show an exposure related increase
in staining of this protein. Staining was equivalent at the 400, 3000,
and 8000 ppm exposures, yet only the two higher concentrations produced
kidney tumors and increased the rate of early mortality.

(iii) Exposure to MTBE or its metabolite TBA increased the
severity of chronic nephropathy in both male and female rats, yet
alpha-2u-globulin is not synthesized in the liver of female rats.
Furthermore, kidney tumors were induced in the mid-dose group of male
rats exposed to TBA, but the severity of nephropathy in that group was
not different from the severity of nephropathy in control male rats.
Thus, other factors must be involved in MTBE- or TBA-induced nephropathy
and renal carcinogenicity in male rats.

Estimations of human cancer risk associated with lifetime
inhalation exposure (70 years) to MTBE are presented in Table 6, using
reasonable worst-case exposure scenarios, estimates of upper-bound unit
cancer risk, and maximum likelihood estimates (MLE) of excess cancer
risk for specific exposures. For lifetime environmental inhalation
exposure to MTBE from use of oxygenated gasoline and reformulated
gasoline containing this ether (based on a reasonable worst-case
exposure scenario of 0.029 ppm), the upper bound excess cancer risk is
2x10-5 using the mouse liver tumor data or the rat kidney tumor
data,
and 1x10-4 using the rat lymphoma/leukemia data. The MLEs of
excess
cancer risk for lifetime environmental exposure to this concentration of
MTBE are 5x10-15 using the mouse liver tumor data,
8x10-10 using the rat
kidney tumor data, and 8x10-5 using the rat lymphoma/leukemia
data. For a working lifetime plus environmental exposure to MTBE among
service station attendants, the estimated upper bound excess cancer
risks were
6x10-5 using the mouse liver tumor data or the rat kidney tumor
data, and 5x10-4 using the rat lymphoma/leukemia data. The
MLEs of excess
cancer risk for service station attendants were 2x10-13 using the
mouse
liver tumor data, 1x10-8 using the rat kidney tumor data, and
3x10-4 using the rat lymphoma/leukemia data. The MLE values are
highly
sensitive to small changes in the experimental tumor incidence data.
For example, if the renal tubule adenoma in the control male rat had
been present instead in the low exposure group, the MLE values based on
the rat kidney tumor data would have been more than 3-4 orders of
magnitude lower than the MLE values shown in Table 6 (see Appendix A).
In this example, the estimated upper-bound excess cancer risk values
change by less than 50%. Because of the instability in the MLEs at low
exposures, they are not considered reliable for estimation of human
risk. While the true risk from low exposures is unknown, the 95% upper
confidence limit provides a stable and plausible upper bounds on the
excess risk.

Sources of uncertainty in these risk assessments include potential
differences in sensitivity between laboratory animals and humans,
interindividual differences in sensitivity among the exposed human
population, the relationship between total dose versus dose rate on the
tumor response, the adequacy of the exposure characterizations
(especially with respect to the distribution of exposures in the
environment and the workplace and the influence of changes in climate),
and the adequacy of the models that were used to perform low-dose
extrapolations and estimate cancer potency. Research that addresses
these uncertainties could either raise or lower the current estimates of
risk.

In estimating excess cancer risks due to exposure to MTBE, several
assumptions were made, including the following:

(i) The annual time-weighted average reasonable worst-case
exposure scenarios used in these calculations and the variabilities in
microenvironmental exposures or in the estimates of duration of time
spent in each activity are within the range of long-term environmental
and occupational lifetime exposures to MTBE among exposed populations.

(ii) Humans and laboratory animals share similar sensitivities to the
carcinogenic effects of MTBE at equivalent doses.

(iii) Tumor data from a gavage study of MTBE can be used to estimate
unit cancer risk for inhalation exposure to MTBE because equivalent
total exposures by inhalation and gavage result in similar internal
doses of MTBE.

(iv) Differences in dose rate and metabolism from gavage and
inhalation exposures do not affect the tumor response.

(v) MTBE-induced nephropathy does not affect the kidney tumor
dose-response.

(vi) At low doses, the carcinogenic effects of MTBE are proportional
to time-weighted average lifetime daily exposures in both experimental
animals and in humans.

(vii) The models used are appropriate for low-dose extrapolations of
the carcinogenic effects of MTBE.

(viii) The carcinogenic effects of MTBE occur independently of the
effects of other hazardous agents to which humans may be exposed. The
effect of addition of MTBE to gasoline on the air concentrations of
other chemical carcinogens that are released as evaporative or
combustion emissions (e.g., benzene, 1,3-butadiene, and formaldehyde) is
beyond the scope of this assessment.

Thus, depending on the validity of the assumptions used in making
these estimates, the actual cancer risks may be lower than those given
in this document and could even be nearly zero. It is not known whether
the cancer risk of oxygenated gasoline containing MTBE is significantly
different from the cancer risk of conventional gasoline. The estimated
upper bound cancer unit risks of MTBE are similar to or slightly lower
than those of fully vaporized conventional gasoline, which has been
listed by EPA as a "probable human carcinogen" based on animal
carcinogenicity data. However, because of a lack of health data on the
nonoxygenated gasoline vapors to which humans are actually exposed, it
is not possible to estimate the population cancer risk to conventional
gasoline. The comparative risk among oxygenated and nonoxygenated
gasoline types has not been established.

The data were generally inadequate to evaluate the health risks of
oxygenates other than MTBE, a factor which makes other oxygenates and
gasoline mixtures to which they are added all the more important to
investigate further.

Note 1. The actual risks are likely to be somewhat lower than the
upper bound calculated risks and could even be nearly zero.Note 2. The MLE values are highly sensitive to small changes in the
tumor incidence data. For example, if the renal tubule adenoma in the
control male rat had been present instead in the low exposure group, the
MLE values based on the rat kidney tumor data would have been more than
3-4 orders of magnitude lower than the MLE values shown in this table
(see Appendix A). In this example, the estimated upper-bound excess
cancer risk values change by less than 50%. Because of their
instability, the MLEs are not considered reliable.

a These cancer risk estimates are derived using a cancer potency
estimate from the gavage carcinogenicity study of MTBE in rats with the
assumption that the observed tumor response in gavage-treated animals
infers an inhalation cancer risk. b Exposure for service station attendants include an 8-hour
time-weighted average occupational exposure (5 days per week for 40
years working lifetime) of 0.6 ppm during the 6-month oxygenated fuel
season and 0.44 ppm during the 6-month reformulated gasoline season.

This assessment has identified several areas where the available
scientific information was considered inadequate to fully assess the
potential health effects, both long-term and short-term, of oxygenated
gasoline. The EPA (1995) has developed a draft of an oxyfuels
information needs report that outlines several research initiatives that
are currently planned or under way to obtain more information about
oxyfuels. Many of the information needs identified in this assessment
are also reflected in the draft EPA report.

In addition, a need exists to more carefully evaluate the
carcinogenicity data on MTBE in terms of its relative potency, as well
as data on subchronic health effects of MTBE and other oxygenates. It
is expected that the more comprehensive review now being undertaken by
the Health Effects Institute will be able to more thoroughly address
these and other complex issues than was possible in this assessment.

Continued use of oxygenated gasoline will require a continuing
evaluation of potential health effects of the oxygenates, their
metabolites, and any emission or atmospheric degradation products. Most
of the research to date has focused on MTBE, and comparatively little
information exists for other oxygenates, such as ETBE or TAME, or some
of the atmospheric degradation products, such as t-butyl formate. In
addition, there is a differential availability of data by compound, with
some endpoints measured for some compounds but not others.

Specifically, further information is needed in the following areas: 1)
more data on human exposures; 2) more data on the pharmacokinetics of
MTBE and its metabolites (TBA and formaldehyde) in humans and animals;
3) epidemiologic and experimental human-subject research on acute health
symptoms related to oxygenates in gasoline mixtures; 4) increased
research on the mechanisms of carcinogenicity and on the dose-response
relationships between exposure to oxygenated fuels and cancer risk; and
5) better data on potential human exposures and health risks associated
with the gasoline mixtures to which oxygenates are added.